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Applied Geochemistry 17 (2002) 517–568
www.elsevier.com/locate/apgeochem
Review
A review of the source, behaviour and distribution
of arsenic in natural waters
P.L. Smedley*, D.G. Kinniburgh
British Geological Survey, Wallingford, Oxon OX10 8BB, UK
Received 1 March 2001; accepted 26 October 2001
Editorial handling by R. Fuge
Abstract
The range of As concentrations found in natural waters is large, ranging from less than 0.5 mg l 1 to more than 5000
mg l 1. Typical concentrations in freshwater are less than 10 mg l 1 and frequently less than 1 mg l 1. Rarely, much
higher concentrations are found, particularly in groundwater. In such areas, more than 10% of wells may be ‘affected’
(defined as those exceeding 50 mg l 1) and in the worst cases, this figure may exceed 90%. Well-known high-As
groundwater areas have been found in Argentina, Chile, Mexico, China and Hungary, and more recently in West
Bengal (India), Bangladesh and Vietnam. The scale of the problem in terms of population exposed to high As concentrations is greatest in the Bengal Basin with more than 40 million people drinking water containing ‘excessive’ As.
These large-scale ‘natural’ As groundwater problem areas tend to be found in two types of environment: firstly, inland
or closed basins in arid or semi-arid areas, and secondly, strongly reducing aquifers often derived from alluvium. Both
environments tend to contain geologically young sediments and to be in flat, low-lying areas where groundwater flow is
sluggish. Historically, these are poorly flushed aquifers and any As released from the sediments following burial has
been able to accumulate in the groundwater. Arsenic-rich groundwaters are also found in geothermal areas and, on a
more localised scale, in areas of mining activity and where oxidation of sulphide minerals has occurred. The As content
of the aquifer materials in major problem aquifers does not appear to be exceptionally high, being normally in the
range 1–20 mg kg 1. There appear to be two distinct ‘triggers’ that can lead to the release of As on a large scale. The
first is the development of high pH (> 8.5) conditions in semi-arid or arid environments usually as a result of the
combined effects of mineral weathering and high evaporation rates. This pH change leads either to the desorption of
adsorbed As (especially As(V) species) and a range of other anion-forming elements (V, B, F, Mo, Se and U) from
mineral oxides, especially Fe oxides, or it prevents them from being adsorbed. The second trigger is the development of
strongly reducing conditions at near-neutral pH values, leading to the desorption of As from mineral oxides and to the
reductive dissolution of Fe and Mn oxides, also leading to As release. Iron (II) and As(III) are relatively abundant in
these groundwaters and SO4 concentrations are small (typically 1 mg l 1 or less). Large concentrations of phosphate,
bicarbonate, silicate and possibly organic matter can enhance the desorption of As because of competition for
adsorption sites. A characteristic feature of high groundwater As areas is the large degree of spatial variability in As
concentrations in the groundwaters. This means that it may be difficult, or impossible, to predict reliably the likely
concentration of As in a particular well from the results of neighbouring wells and means that there is little alternative
but to analyse each well. Arsenic-affected aquifers are restricted to certain environments and appear to be the exception
rather than the rule. In most aquifers, the majority of wells are likely to be unaffected, even when, for example, they
contain high concentrations of dissolved Fe. # 2002 Published by Elsevier Science Ltd. All rights reserved.
* Corresponding author. Fax: +44-1491-692345.
E-mail address: [email protected] (P.L. Smedley).
0883-2927/02/$ - see front matter # 2002 Published by Elsevier Science Ltd. All rights reserved.
PII: S0883-2927(02)00018-5
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P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
Contents
1. Introduction ........................................................................................................................................................... 519
2. Arsenic in natural waters ....................................................................................................................................... 520
2.1. Aqueous speciation........................................................................................................................................ 520
2.2. Abundance and distribution.......................................................................................................................... 520
2.2.1. Atmospheric precipitation ................................................................................................................. 521
2.2.2. River water ........................................................................................................................................ 523
2.2.3. Lake water......................................................................................................................................... 524
2.2.4. Seawater and estuaries....................................................................................................................... 525
2.2.5. Groundwater ..................................................................................................................................... 525
2.2.6. Mine drainage.................................................................................................................................... 525
2.2.7. Sediment porewaters.......................................................................................................................... 525
2.2.8. Oilfield and other brines .................................................................................................................... 526
2.3. Distribution of arsenic species in water bodies.............................................................................................. 526
2.4. Impact of redox kinetics on arsenic speciation.............................................................................................. 527
3. Sources of arsenic................................................................................................................................................... 528
3.1. Minerals......................................................................................................................................................... 528
3.1.1. Major arsenic minerals ...................................................................................................................... 528
3.1.2. Rock-forming minerals...................................................................................................................... 529
3.2. Rocks, sediments and soils ............................................................................................................................ 530
3.2.1. Igneous rocks..................................................................................................................................... 530
3.2.2. Metamorphic rocks ........................................................................................................................... 530
3.2.3. Sedimentary rocks ............................................................................................................................. 530
3.2.4. Unconsolidated sediments ................................................................................................................. 532
3.2.5. Soils ................................................................................................................................................... 533
3.2.6. Contaminated surficial deposits......................................................................................................... 533
3.3. The atmosphere ............................................................................................................................................. 533
4. Mineral-water interactions ..................................................................................................................................... 533
4.1. Controls on arsenic mobilisation................................................................................................................... 533
4.2. Arsenic associations in sediments .................................................................................................................. 534
4.3. Reduced sediments and the role of iron oxides ............................................................................................. 534
4.4. Arsenic release from soils and sediments following reduction....................................................................... 537
4.5. Speciation of elements in sediments and the role of selective extraction techniques ..................................... 538
4.6. Transport of arsenic ...................................................................................................................................... 539
5. Groundwater environments with high arsenic concentrations ............................................................................... 542
5.1. World distribution of groundwater arsenic problems ................................................................................... 542
5.2. Reducing environments ................................................................................................................................. 543
5.2.1. Bangladesh and India (West Bengal)................................................................................................. 543
5.2.2. Taiwan............................................................................................................................................... 547
5.2.3. Northern china .................................................................................................................................. 547
5.2.4. Vietnam ............................................................................................................................................. 547
5.2.5. Hungary and Romania...................................................................................................................... 548
5.3. Arid oxidising environments.......................................................................................................................... 548
5.3.1. Mexico............................................................................................................................................... 548
5.3.2. Chile .................................................................................................................................................. 548
5.3.3. Argentina........................................................................................................................................... 549
5.4. Mixed oxidising and reducing environments ................................................................................................. 549
5.4.1. South-western USA ........................................................................................................................... 549
5.5. Geothermal sources ....................................................................................................................................... 550
5.6. Sulphide mineralisation and mining-related arsenic problems ...................................................................... 551
5.6.1. Thailand ............................................................................................................................................ 551
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
519
5.6.2. Ghana................................................................................................................................................ 551
5.6.3. United States ..................................................................................................................................... 551
5.6.4. Other areas ........................................................................................................................................ 551
6. Common features of groundwater arsenic problem areas...................................................................................... 552
6.1. A hydrogeochemical perspective ................................................................................................................... 552
6.2. The source of arsenic..................................................................................................................................... 552
6.3. Arsenic mobilisation—the necessary geochemical trigger ............................................................................. 552
6.3.1. Desorption at high pH under oxidising conditions ........................................................................... 553
6.3.2. Arsenic desorption and dissolution due to a change to reducing conditions .................................... 554
6.3.3. Reduction in surface area of oxide minerals ..................................................................................... 555
6.3.4. Reduction in binding strength between arsenic and mineral surfaces ............................................... 555
6.3.5. Mineral dissolution............................................................................................................................ 556
6.4. Transport—historical groundwater flows...................................................................................................... 556
6.5. Future developments in arsenic research....................................................................................................... 558
6.6. Identification of ‘at risk’ aquifers .................................................................................................................. 558
7. Concluding remarks ............................................................................................................................................... 559
Acknowledgements...................................................................................................................................................... 560
References ................................................................................................................................................................... 560
1. Introduction
The recent finding that groundwaters from large areas
of West Bengal, Bangladesh and elsewhere are heavily
enriched with As has prompted a reassessment of the
factors controlling the distribution of As in the natural
environment and the ways in which As may be mobilised.
Arsenic is a ubiquitous element found in the atmosphere,
soils and rocks, natural waters and organisms. It is mobilised through a combination of natural processes such as
weathering reactions, biological activity and volcanic
emissions as well as through a range of anthropogenic
activities. Most environmental As problems are the result
of mobilisation under natural conditions. However,
man has had an important additional impact through
mining activity, combustion of fossil fuels, the use of
arsenical pesticides, herbicides and crop desiccants and
the use of As as an additive to livestock feed, particularly
for poultry. Although the use of arsenical products such
as pesticides and herbicides has decreased significantly in
the last few decades, their use for wood preservation is
still common. The impact on the environment of the use
of arsenical compounds, at least locally, will remain for
some years.
Of the various sources of As in the environment,
drinking water probably poses the greatest threat to
human health. Airborne As, particularly through occupational exposure, has also given rise to known health
problems in some areas. Drinking water is derived from
a variety of sources depending on local availability: surface water (rivers, lakes, reservoirs and ponds), groundwater (aquifers) and rain water. These sources are very
variable in terms of As risk. Alongside obvious point
sources of As contamination, high concentrations are
mainly found in groundwaters. These are where the
greatest number of, as yet unidentified, sources are likely
to be found. This review therefore focuses on the factors
controlling As concentrations in groundwaters. However, the authors also review the occurrence of As in a
broad range of natural waters since these may indirectly
be involved in the formation of As-rich groundwaters and
can also provide a useful background against which to
view groundwater As concentrations. Furthermore, many
of the processes involved in the uptake and release of As
are common to a wide range of natural environments.
Following the accumulation of evidence for the
chronic toxicological effects of As in drinking water,
recommended and regulatory limits of many authorities
are being reduced. The WHO guideline value for As in
drinking water was provisionally reduced in 1993 from
50 to 10 mg l 1. The new recommended value was based
on the increasing awareness of the toxicity of As, particularly its carcinogenicity, and on the ability to measure
it quantitatively (WHO, 1993). If the standard basis for
risk assessment applied to industrial chemicals were
applied to As, the maximum permissible concentration
would be lower still. The EC maximum admissible concentration (MAC) for As in drinking water has been
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P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
reduced to 10 mg l 1. The Japanese limit for drinking
water is also 10 mg l 1 while the interim maximum
acceptable concentration for Canadian drinking water is
25 mg l 1. The US-EPA limit was also reduced from 50
to 10 mg l 1 in January 2001 following prolonged debate
over the most appropriate limit. However, this rule is
now (September 2001) being reconsidered given the high
cost implications to the US water industry, estimated at
$200 million per year. Whilst many national authorities
are seeking to reduce their limits in line with the WHO
guideline value, many countries and indeed all affected
developing countries, still operate at present to the 50 mg
l 1 standard, in part because of lack of adequate testing
facilities for lower concentrations.
Until recently, As was often not on the list of constituents in drinking water routinely analysed by
national laboratories, water utilities and non-governmental organizations (NGOs) and so the body of information about the distribution of As in drinking water is
not as well known as for many other drinking-water
constituents. In recent years, it has become apparent
that both the WHO guideline value and current national
standards are quite frequently exceeded in drinkingwater sources, and often unexpectedly so. Indeed, As and
F are now recognised as the most serious inorganic contaminants in drinking water on a worldwide basis. In
areas of high As concentrations, drinking water provides
a potentially major source of As in the diet and so its
early detection is of considerable importance.
2. Arsenic in natural waters
2.1. Aqueous speciation
Arsenic is perhaps unique among the heavy metalloids and oxyanion-forming elements (e.g. As, Se, Sb, Mo,
V, Cr, U, Re) in its sensitivity to mobilisation at the pH
values typically found in groundwaters (pH 6.5–8.5) and
under both oxidising and reducing conditions. Arsenic can
occur in the environment in several oxidation states ( 3,
0, +3 and +5) but in natural waters is mostly found in
inorganic form as oxyanions of trivalent arsenite
[As(III)] or pentavalent arsenate [As(V)]. Organic As
forms may be produced by biological activity, mostly in
surface waters, but are rarely quantitatively important.
Organic forms may however occur where waters are
significantly impacted by industrial pollution.
Most toxic trace metals occur in solution as cations
(e.g. Pb2+, Cu2+, Ni2+, Cd2+, Co2+, Zn2+) which generally become increasingly insoluble as the pH increases.
At the near-neutral pH typical of most groundwaters, the
solubility of most trace-metal cations is severely limited
by precipitation as, or coprecipitation with, an oxide,
hydroxide, carbonate or phosphate mineral, or more
likely by their strong adsorption to hydrous metal oxides,
clay or organic matter. In contrast, most oxyanions
including arsenate tend to become less strongly sorbed
as the pH increases (Dzombak and Morel, 1990). Under
some conditions at least, these anions can persist in
solution at relatively high concentrations (tens of mg l 1)
even at near-neutral pH values. Therefore the oxyanionforming elements such as Cr, As, U and Se are some of
the most common trace contaminants in groundwaters.
Relative to the other oxyanion-forming elements, As is
among the most problematic in the environment because
of its relative mobility over a wide range of redox conditions. Selenium is mobile as the selenate (SeO24 ) oxyanion
under oxidising conditions but is immobilized under
reducing conditions either due to the stronger adsorption of its reduced form, selenite (SeO23 ), or due to its
reduction to the metal. Chromium can similarly be
mobilized as stable Cr(VI) oxyanion species under oxidising conditions, but forms cationic Cr(III) species in
reducing environments and hence behaves like other trace
cations (i.e. is relatively immobile at near-neutral pH
values). Other oxyanions such as molybdate, vanadate,
uranyl and rhenate also appear to be less mobile under
reducing conditions. In S-rich, reducing environments,
many of the trace metals also form insoluble sulphides.
Arsenic is distinctive in being relatively mobile under
reduced conditions. It can be found at concentrations in
the mg l 1 range when all other oxyanion-forming
elements are present in the mg l 1 range.
Redox potential (Eh) and pH are the most important
factors controlling As speciation. Under oxidising conditions, H2AsO4 is dominant at low pH (less than about
pH 6.9), whilst at higher pH, HAsO24 becomes dominant (H3AsO04 and AsO34 may be present in extremely
acidic and alkaline conditions respectively). Under
reducing conditions at pH less than about pH 9.2, the
uncharged arsenite species H3AsO03 will predominate
(Fig. 1; Brookins, 1988; Yan et al., 2000). The distributions of the species as a function of pH are given in
Fig. 2. In practice, most studies in the literature report
speciation data without consideration of the degree of
protonation. In the presence of extremely high concentrations of reduced S, dissolved As-sulphide species
can be significant. Reducing, acidic conditions favour
precipitation of orpiment (As2S3), realgar (AsS) or other
sulphide minerals containing coprecipitated As (Cullen
and Reimer, 1989). Therefore high-As waters are not
expected where there is a high concentration of free
sulphide (Moore et al., 1988).
2.2. Abundance and distribution
Concentrations of As in fresh water vary by more
than four orders of magnitude (Table 1) depending on the
source of As, the amount available and the local geochemical environment. Under natural conditions, the
greatest range and the highest concentrations of As are
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521
Fig. 1. Eh-pH diagram for aqueous As species in the system
As–O2–H2O at 25 C and 1 bar total pressure.
found in groundwaters as a result of the strong influence
of water-rock interactions and the greater tendency in
aquifers for the physical and geochemical conditions to
be favourable for As mobilization and accumulation.
The range of concentrations for many water bodies is
large and hence ‘typical’ values are difficult to derive.
Many studies of As reported in the literature have also
preferentially targeted known problem areas and hence
reported ranges are often extreme and unrepresentative
of natural waters as a whole. Nonetheless, the following
compilation of data for ranges of As concentrations found
in various parts of the hydrosphere and lithosphere gives a
broad indication of the expected concentration ranges and
their variation in the environment.
2.2.1. Atmospheric precipitation
Arsenic enters the atmosphere through inputs from
wind erosion, volcanic emissions, low-temperature
volatilisation from soils, marine aerosols and pollution
and is returned to the earth’s surface by wet and dry
deposition. The most important anthropogenic inputs
are from smelter operations and fossil-fuel combustion.
The As appears to consist of mainly As(III)2O3 dust
particles (Cullen and Reimer, 1989). Nriagu and Pacyna
(1988) estimated that anthropogenic sources of atmospheric arsenic (around 18,800 tonnes a 1) amounted to
around 70% of the global atmospheric As flux. While it
is accepted that these anthropogenic sources have an
important impact on airborne As compositions, their
influence on the overall As cycle is not well established.
Baseline concentrations of As in rainfall and snow in
rural areas are invariably low at typically less than 0.03
mg l 1 (Table 1). Concentrations in areas affected by
Fig. 2. (a) Arsenite and (b) arsenate speciation as a function of
pH (ionic strength of about 0.01 M). Redox conditions have
been chosen such that the indicated oxidation state dominates
the speciation in both cases.
smelter operations, coal burning and volcanic emissions
are generally higher. Andreae (1980) found rainfall
potentially affected by smelting and coal burning to
have As concentrations of around 0.5 mg l 1 (Table 1),
although higher concentrations (average 16 mg l 1) have
been found in rainfall collected in Seattle some 35 km
downwind of a Cu smelter (Crecelius, 1975). Values
given for Arizona snowpacks (Table 1; Barbaris and
Betterton, 1996) are also probably slightly above baseline concentrations because of potential inputs of airborne As from smelters, power plants and soil dust. In
general however, sources of airborne As in most industrialized nations are limited as a result of air-pollution
control measures. Unless significantly contaminated
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Table 1
Typical As concentrations in natural waters
As concentration average
or range (mg l 1)
Reference
Rain water
Baseline
Maritime
Terrestrial (w USA)
Coastal (Mid-Atlantic, USA)
Snow (Arizona)
0.02
0.013–0.032
0.1 (< 0.005–1.1)
0.14 (0.02–0.42)
Andreae (1980)
Andreae (1980)
Scudlark and Church (1988)
Barbaris and Betterton (1996)
Non-baseline:
Terrestrial rain
Seattle rain, impacted by copper smelter
0.46
16
Andreae (1980)
Crecelius (1975)
River water
Baseline
Various
0.83 (0.13–2.1)
0.25 ( <0.02–1.1)
0.15–0.45
2.1
0.7
1.3
4.5–45
3 (1–8)
0.75–3.8 (up to 30)
Andreae et al. (1983); Froelich et al. (1985);
Seyler and Martin (1991)
Lenvik et al. (1978)
Waslenchuk (1979)
Sonderegger and Ohguchi (1988)
Seyler and Martin (1990)
Pettine et al. (1992)
Seyler and Martin (1990)
Quentin and Winkler (1974)
Andreae and Andreae (1989)
190–21800
400–450
7–114
Cáceres et al. (1992)
Sancha (1999)
Lerda and Prosperi (1996)
0.20–264
32 (28–36)
44 (19–67)
10–370
Benson and Spencer (1983)
McLaren and Kim (1995)
Robinson et al. (1995)
Nimick et al. (1998)
Mining influenced
Ron Phibun, Thailand
Ashanti, Ghana
British Columbia, Canada
218 (4.8–583)
284 ( <2–7900)
17.5 ( <0.2–556)
Williams et al. (1996)
Smedley et al. (1996)
Azcue et al. (1994)
Lake water
Baseline
British Columbia
Ontario
France
Japan
Sweden
0.28 ( <0.2–0.42)
0.7
0.73–9.2 (high Fe)
0.38–1.9
0.06–1.2
Azcue et al. (1994, 1995)
Azcue and Nriagu (1995)
Seyler and Martin (1989)
Baur and Onishi (1969)
Reuther (1992)
Geothermal influenced
Western USA
0.38–1000
Benson and Spencer (1983)
Mining influenced
Northwest Territories, Canada
Ontario, Canada
270 (64–530)
35–100
Bright et al. (1996)
Azcue and Nriagu (1995)
Estuarine water
Baseline
Oslofjord, Norway
Saanich Inlet, British Columbia
0.7–2.0
1.2–2.5
Abdullah et al. (1995)
Peterson and Carpenter (1983)
Water body and location
Norway
South-east USA
USA
Dordogne, France
Po River, Italy
Polluted European rivers
River Danube, Bavaria
Schelde catchment, Belgium
High-As groundwater influenced:
Northern Chile
Northern Chile
Córdoba, Argentina
Geothermal influenced
Sierra Nevada, USA
Waikato, New Zealand
Madison and Missouri Rivers, USA
(continued on next page)
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523
Table 1 (continued)
Water body and location
As concentration average
or range (mg l 1)
Reference
Rhône Estuary, France
Krka Estuary, Yugoslavia
2.2 (1.1–3.8)
0.13–1.8
Seyler and Martin (1990)
Seyler and Martin (1991)
Mining and industry influenced
Loire Estuary, France
Tamar Estuary, UK
Schelde Estuary, Belgium
up to 16
2.7–8.8
1.8–4.9
Seyler and Martin (1990)
Howard et al. (1988)
Andreae and Andreae (1989)
Seawater
Deep Pacific and Atlantic
Coastal Malaysia
Coastal Spain
Coastal Australia
1.0–1.8
1.0 (0.7–1.8)
1.5 (0.5–3.7)
1.3 (1.1–1.6)
Cullen and Reimer (1989)
Yusof et al. (1994)
Navarro et al. (1993)
Maher (1985)
< 0.5–10
10–5000
Groundwater
Baseline UK
As-rich provinces (e.g. Bengal
Basin, Argentina, Mexico, northern
China, Taiwan, Hungary)
Mining-contaminated groundwaters
50–10,000
Geothermal water
< 10–50,000
Arsenical herbicide plant, Texas
408,000
Edmunds et al. (1989)
Das et al. (1995); BGS and DPHE (2001);
Nicolli et al. (1989); Smedley et al. (2001a);
Del Razo et al. (1990); Luo et al. (1997);
Hsu et al. (1997); Varsányi et al. (1991)
Wilson and Hawkins (1978);Welch et al. (1988);
Williams et al. (1996)
Baur and Onishi (1969); White et al., (1963),
Ellis and Mahon (1977)
Kuhlmeier (1997a,b)
Mine drainage
Various, USA
Iron Mountain
Ural Mountains
< 1–34,000
up to 850,000
400,000
Plumlee et al. (1999)
Nordstrom and Alpers (1999)
Gelova (1977)
1.3–166
3.2–99
Widerlund and Ingri (1995)
Yan et al. (2000)
up to 300
50–360
300–100,000
Sullivan and Aller (1996)
Azcue et al. (1994)
McCreadie et al. (2000)
230
up to 243,000
White et al. (1963)
White et al. (1963)
Sediment porewater
Baseline, Swedish Estuary
Baseline, clays, Saskatchewan,
Canada
Baseline, Amazon shelf sediments
Mining-contam’d, British Columbia
Tailings impoundment, Ontario,
Canada
Oilfield and related brine
Ellis Pool, Alberta, Canada
Searles Lake brine, California
with industrial sources of As, atmospheric precipitation
contributes little As to surface and groundwater bodies.
2.2.2. River water
Baseline concentrations of As in river waters are also
low (in the region of 0.1–0.8 mg l 1 but can range up to
ca. 2 mg l 1; Table 1). They vary according to the composition of the surface recharge, the contribution from
baseflow and the bedrock lithology. Concentrations at
the low end of the range have been found in rivers
draining As-poor bedrocks. Seyler and Martin (1991)
found average river concentrations as low as 0.13 mg l 1
in the Krka region of Yugoslavia where the bedrock is
As-poor karstic limestone (Table 1). Lenvik et al. (1978)
also found low average concentrations of about 0.25 mg
l 1 in rivers draining basement rocks in Norway, the
lowest being in catchments on Precambrian rocks.
Waslenchuk (1979) found concentrations in river waters
from the south-eastern USA in the range 0.15–0.45 mg
l 1 (Table 1).
Relatively high concentrations of naturally-occurring
As can occur in some areas as a result of inputs from
geothermal sources or high-As groundwaters. Arsenic
concentrations in river waters from geothermal areas
have been reported typically at around 10–70 mg l 1
(e.g. western USA and New Zealand; McLaren and
Kim, 1995; Robinson et al., 1995; Nimick et al., 1998;
Table 1), although higher concentrations have been
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found. Nimick et al. (1998) for example found As concentrations up to 370 mg l 1 in Madison River water
(Wyoming and Montana) as a result of geothermal
inputs from the Yellowstone geothermal system. Wilkie
and Hering (1998) also found concentrations in the
range 85–153 mg l 1 in Hot Creek (tributary of the
Owens River, California). Some river waters affected by
geothermal activity show distinct seasonal variations in
As concentration. Concentrations in the Madison River
have been noted to be highest during low-flow conditions. This has been attributed to a greater contribution
of geothermal water during times of low flow and dilution from spring runoff at times of high flow (Nimick et
al., 1998). In the Waikato river system of New Zealand,
As maxima were found in the summer months. These
increases were linked to temperature-controlled microbial reduction of As(V) to As(III) with consequent
increased mobility of As(III) (McLaren and Kim, 1995).
Increased concentrations are also reported in some
river waters from arid areas where the surface water is
dominated by river baseflow. The resulting surface
waters often have a high pH and alkalinity. For example,
in surface waters from the Loa River Basin of northern
Chile (Antofagasta area, Atacama desert), Cáceres et al.
(1992) found concentrations of naturally-occurring As
ranging between 190 and 21,800 mg l 1. The high As
concentrations correlated well with salinity. While geothermal inputs are likely to have had an important
impact on the chemical compositions of the river waters
in this area (Section 5.5), evaporative concentration of
baseflow-dominated river water is also likely to be
important in the arid conditions. Increased As concentrations (up to 114 mg l 1) have also been reported in
river waters from central Argentina where regional
groundwater-As concentrations (and pH, alkalinity) are
high (Lerda and Prosperi, 1996).
Although bedrock inevitably has an influence on
river-water As concentrations, concentrations in rivers
with more typical pH and alkalinity values (c. pH 5–7,
alkalinity <100 mg l 1 as HCO3) do not show the extremely high concentrations found in groundwaters because
of oxidation and adsorption of As species onto the river
sediments as well as dilution by surface recharge and runoff. Arsenic concentrations in seven river water samples
from Bangladesh have been reported in the range < 0.5–2.7
mg l 1 but with one sample having a high concentration of
29 mg l 1 (BGS and DPHE, 2001). The highest value
observed is significantly above world-average baseline
concentrations (Table 1) but is much lower than some of
the values found in the groundwaters.
Significant increases in As concentrations of river
waters may also occur as a result of pollution from
industrial or sewage effluents. Andreae and Andreae
(1989) found concentrations up to 30 mg l 1 in water
from the River Zenne, Belgium which is affected by
inputs from urban and industrial sources, particularly
sewage. However, the concentration of As in water from
most of the catchment was in the range 0.75–3.8 mg l 1
and not significantly different from baseline concentrations. Durum et al. (1971) reported As concentrations in
727 samples of surface waters from the United States.
While 79% of the samples had As concentrations below
the (rather high) detection limit of 10 mg l 1, the highest
observed concentration, 1100 mg l 1, was found in Sugar
Creek, South Carolina, downstream of an industrial
complex.
Arsenic can also be derived from mine wastes and mill
tailings. Azcue and Nriagu (1995) found baseline concentrations in the Moira River, Ontario of 0.7 mg l 1
upstream of the influence of tailings from gold-mine
workings. Downstream, concentrations increased to 23
mg l 1. Azcue et al. (1994) found concentrations up to
556 mg l 1 (average 17.5 mg l 1) in streams adjacent to
tailings deposits in British Columbia. Williams et al.
(1996) and Smedley et al. (1996) noted high As concentrations (typically around 200–300 mg l 1) in surface
waters affected respectively by Sn- and Au-mining activities. Though often involving notable increases above
baseline concentrations, such anomalies tend to be relatively localised around the pollution source, principally
because of the strong adsorption affinity of oxide minerals, especially Fe oxide, for As under oxidising, neutral
to mildly acidic conditions.
2.2.3. Lake water
Concentrations of As in lake waters are typically close
to or lower than those found in river water. Baseline
concentrations have been found at < 1 mg l 1 in Canada
(Azcue and Nriagu, 1995; Azcue et al., 1995). As with
river waters, increased concentrations are found in lake
waters affected by geothermal water and by mining
activity. Ranges of typically 100–500 mg l 1 have been
reported in some mining areas and up to 1000 mg l 1 in
geothermal areas (Table 1). Arsenic concentrations in
mining-affected lake waters are not always high however, as removal from solution can be achieved effectively by adsorption onto Fe oxides under neutral to
mildly acidic conditions. Azcue et al. (1994), for example, found As concentrations in Canadian lake waters
affected by mining effluent similar to those not affected
by mining effluent, in each case about 0.3 mg l 1.
High As concentrations are also found in some alkaline closed-basin lakes as a result of extreme evaporation and/or geothermal inputs. Mono Lake in the
California, USA, for example, has concentrations of
dissolved As of 10,000–20,000 mg l 1, with pH values in
the range 9.5–10 as a result of inputs from geothermal
springs and the weathering of volcanic rocks followed
by evaporation (Maest et al., 1992).
There is also much evidence for stratification of As
concentrations in some lake waters as a result of varying
redox conditions (Aggett and O’Brien, 1985). Azcue and
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
Nriagu (1995) found that concentrations increased with
depth (up to 10 m) in lake waters from Ontario, probably because of an increasing ratio of As(III) to As(V)
with depth and an influx of mining-contaminated sediment porewaters at the sediment-water interface. The
concentrations were higher in summer when the proportion of As(III) was observed to be higher. Depleted
O2 levels in the bottom lake waters as a result of biological productivity during the summer months are a
likely cause of the higher As concentrations in the deeper
lake waters.
2.2.4. Seawater and estuaries
Average As concentrations in open seawater usually
show little variation and are typically around 1.5 mg l 1
(Table 1). Concentrations in estuarine water are more
variable as a result of varying river inputs and salinity or
redox gradients but are also usually low, at typically less
than 4 mg l 1 under natural conditions. Peterson and
Carpenter (1983) found concentrations between 1.2–2.5
mg l 1 in waters from Saanich Inlet, British Columbia.
Values less than 2 mg l 1 were found in Oslofjord, Norway (Abdullah et al., 1995; Table 1). Concentrations are
commonly higher when riverine inputs are affected by
industrial or mining effluent (e.g. Tamar, Schelde, Loire
Estuaries; Table 1) or by geothermal water. Unlike some
other trace elements such as B, saline intrusion of seawater into an aquifer is unlikely to lead to a significant
increase of As in the affected groundwater.
Arsenate shares many chemical characteristics with
phosphate and hence in oxic marine and estuarine waters,
depletions in phosphate in biologically productive surface
waters are mirrored by depletions in arsenate. Arsenate
concentration minima often coincide with photosynthetic
maxima evidenced by high concentrations of chlorophyll a (Cullen and Reimer, 1989). Several studies have
noted variations in the behaviour of As during estuarine
mixing. Some have reported conservative behaviour. In
the unpolluted Krka Estuary of Yugoslavia, Seyler and
Martin (1991) observed a linear increase in total As with
increasing salinity ranging from 0.13 mg l 1 in fresh
waters to 1.8 mg l 1 offshore (i.e. seawater value). However, other studies have observed non-conservative
behaviour (departures from simple mixing) in estuaries
due to processes such as diffusion from sediment porewaters, coprecipitation with Fe oxides or anthropogenic
inputs (e.g. Andreae et al., 1983; Andreae and Andreae,
1989). The flocculation of Fe oxides at the freshwatersaline interface is an important consequence of increases
in pH and salinity. This can lead to major decreases in
the As flux to the oceans (Cullen and Reimer, 1989).
2.2.5. Groundwater
Background concentrations of As in groundwater are
in most countries less than 10 mg l 1 (e.g. Edmunds et
al., 1989 for the UK; Welch et al., 2000 for the USA)
525
and sometimes substantially lower. However, values
quoted in the literature show a very large range from
< 0.5 to 5000 mg l 1 (i.e. four orders of magnitude). This
range occurs under natural conditions. High concentrations of As are found in groundwater in a variety of
environments. These include both oxidising (under conditions of high pH) and reducing aquifers and areas affected
by geothermal, mining and industrial activity. Evaporative concentration can also increase concentrations substantially. Most high-As groundwater provinces are the
result of natural occurrences of As. Cases of mininginduced As pollution are numerous in the literature but
tend to be localised. Cases of industrially-induced As
pollution (including that from agriculture) may be
severe locally (Table 1) but occurrences are relatively
rare. Arsenic occurrences in groundwater are discussed
more fully in Section 5.
2.2.6. Mine drainage
Under the extremely acid conditions of some acid
mine drainage, which can have negative pH values
(Nordstrom et al., 2000), high concentrations of a wide
range of solutes are found, including Fe and As. The
highest reported As concentration of 850,000 mg l 1 is
from an acid seep in the Richmond mine at Iron
Mountain, California (Nordstrom and Alpers, 1999). In
a compilation of some 180 samples of mine drainage
from the USA, Plumlee et al. (1999) reported concentrations ranging from detection limits ( <1 mg l 1 or more)
to 340,000 mg l 1, again the highest values being from
the Richmond mine. Gelova (1977) also reported an As
concentration of 400,000 mg l 1 from the Ural Mountains. Dissolved As in acid mine waters is rapidly
removed as the Fe is oxidised and precipitated and the
As scavenged through adsorption. At Iron Mountain,
an efficient neutralization plant removes the As and
metals for safe disposal.
2.2.7. Sediment porewaters
Some high concentrations of As have been found in
porewaters extracted from unconsolidated sediments
and often form sharp contrasts to the concentrations
observed in overlying surface waters (e.g. Belzile and
Tessier, 1990). Widerlund and Ingri (1995) found concentrations in the range 1.3–166 mg l 1 in porewaters
from the Kalix River estuary of northern Sweden. Yan
et al. (2000) found As concentrations in the range 3.2–99
mg l 1 in porewaters from clay sediments in Saskatchewan, Canada (Table 1). Increased concentrations have
been found in porewaters affected by geothermal inputs.
Aggett and Kriegman (1988) found As concentrations up
to 6430 mg l 1 in anoxic porewaters from New Zealand.
Even higher concentrations can be found in porewaters
from sediments affected by mining contamination (tailings, mineral-rich deposits). McCreadie et al. (2000)
reported As concentrations up to 100,000 mg l 1 in
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porewaters extracted from tailings in Ontario (Table 1).
In such cases, high porewater As concentrations are
most likely to be linked to the strong redox gradients
that occur below the sediment-water interface, often
over depth scales of centimeters. Burial of fresh organic
matter and the slow diffusion of O2 through the sediment leads to reducing conditions just below the sediment-water interface. This encourages the reduction of
As(V) and desorption from Fe and Mn oxides, as well as
reductive dissolution of these minerals. There is much
evidence for cycling of As between shallow sediment
porewaters and overlying surface waters in response to
temporal variations in redox conditions.
Sullivan and Aller (1996) carried out an elegant study
of the cycling of As in shallow sediments from the offshore shelf of the Amazon situated far from population
centres. They measured porewater As and Fe concentration profiles as well as sediment As and Fe(II)
concentrations. There was frequently a well-correlated
peak in dissolved As and Fe concentrations some 50–
150 cm beneath the surface, with As concentrations in
the peak averaging about 135 mg l 1 and reaching a
maximum of 300 mg l 1, much greater than from marine
coastal environments. The dissolved As/Fe molar ratio
varied but was typically about 1:300. Dissolved As varied inversely with easily-leachable (6 M HCl) As in the
sediment and increased directly with solid-phase Fe(II).
In these sediments, Fe oxides were believed to be a
much more important source of As than Mn oxides.
2.2.8. Oilfield and other brines
Only limited data are available for As in oilfield and
other brines, but some published accounts suggest that
concentrations can be very high. White et al. (1963)
reported a dissolved As concentration of 230 mg l 1 in a
Na–HCO3 groundwater from a 1000 m deep oilfield well
from Ellis Pool, Alberta, Canada. They also reported a
concentration of 5800 mg l 1 As in a Na–Cl-dominated
brine from Tisakürt, Hungary. Composite brines from
the interstices of salt deposits from Searles Lake, California, have As concentrations up to 243 mg l 1 (Na 119
g l 1; White et al., 1963; Table 1).
2.3. Distribution of arsenic species in water bodies
Many studies of As speciation in natural waters have
been carried out. Most attempt to separate the inorganic
species into As(III) and As(V), usually by chromatographic separation or by making use of the relatively
slow reduction of As(V) by Na borohydride. Some studies
also measure the organic As species. The sampling and
analytical techniques required are not trivial and not yet
well-established (Edwards et al., 1998). Both oxidation
of As(III) and reduction of As(V) may occur during storage (Hall et al., 1999). Separation of species may be carried out in the field to avoid the problem of preserving
species for later laboratory analysis. Alternatively, preservation with HCl and ascorbic acid has been successful although this may destroy monomethylarsonic acid
(MMAA) if present.
In rain water, oxidation states of As present will vary
according to the source. This is likely to be dominantly
As(III)2O3 when derived from smelters, coal burning
and volcanic sources, although organic species may be
derived by volatilization from soils, arsine (As(-III)H3)
may derive from landfills and reducing soils such as
peats, and arsenate may be derived from marine aerosols. Reduced forms will undergo oxidation by O2 in the
atmosphere and reactions with atmospheric SO2 or O3
are likely (Cullen and Reimer, 1989).
In oxic seawater, the As is typically dominated by
As(V), though some As(III) is invariably present and
becomes of increasing importance in anoxic bottom
waters. Ratios of As(V)/As(III) are typically in the
range 10–100 in open seawater (Andreae, 1979; Peterson
and Carpenter, 1983; Pettine et al., 1992). Arsenic(V)
should exist mainly as HAsO24 and H2AsO4 in the pH
range of seawater (pH around 8.2; Figs. 1 and 2) and
As(III) mainly as the neutral species H3AsO3. Relatively
high proportions of H3AsO3 are found in surface ocean
waters (Cullen and Reimer, 1989; Cutter et al., 2001).
These coincide with zones of primary productivity.
Increases in organic As species have also been recorded
in these zones as a result of methylation reactions by
phytoplankton.
The relative proportions of As species are more variable
in estuarine waters because of variable redox and salinity,
and terrestrial inputs (Howard et al., 1988; Abdullah et al.,
1995). However, they are still dominated by As(V).
Andreae and Andreae (1989) found As(V)/As(III) ratios
varying between 5–50 in the Schelde Estuary of Belgium
with the lowest ratios in anoxic zones where inputs of
industrial effluent had an impact. Increased proportions
of As(III) also result from inputs of mine effluent
(Klumpp and Peterson, 1979). Seasonal variations in As
concentration and speciation have been noted in estuaries (Riedel, 1993). In seasonally anoxic estuarine
waters, variations in the relative proportions of As(III)
and As(V) can be large. Peterson and Carpenter (1983)
found a distinct crossover in the proportions of the two
species with increasing depth in response to the onset of
anoxic conditions in the estuarine waters of Saanich
Inlet of British Columbia. Arsenic(III) represented only
5% (0.10 mg l 1) of the dissolved As above the redox
front but 87% (1.58 mg l 1) below it. In marine and
estuarine waters, organic forms are usually less abundant but are nonetheless often detected (e.g. Riedel,
1993; Howard et al., 1999). Concentrations of these will
depend on abundance and species of biota present and
on temperature.
In lake and river waters, As(V) is also generally the
dominant species (e.g. Seyler and Martin, 1990; Pettine
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
et al., 1992), though significant seasonal variations in
speciation as well as absolute concentration have been
found. Concentrations and relative proportions of As(V)
and As(III) vary according to changes in input sources,
redox conditions and biological activity. The presence of
As(III) may be maintained in oxic waters by biological
reduction of As(V), particularly during summer months.
Higher relative proportions of As(III) have been found
in river stretches close to inputs of As(III)-dominated
industrial effluent (Andreae and Andreae, 1989) and in
waters with a component of geothermal water.
Proportions of As(III) and As(V) are particularly
variable in stratified lakes where redox gradients can be
large and seasonally variable (Kuhn and Sigg, 1993). As
with estuarine waters, distinct changes in As speciation
occur in lake profiles as a result of redox changes. For
example, in the stratified, hypersaline and hyperalkaline
Mono Lake (California, USA), there is a predominance
of As(V) in the upper oxic layer and of As(III) in the
reducing layer (Maest et al., 1992; Oremland et al.,
2000). Rapid oxidation of As(III) occurs as a result of
microbial activity during the early stages of lake turnover (Oremland et al., 2000). The As oxidation occurs
before Fe(II) oxidation. Unlike Mono Lake, speciation
of As in lakes does not necessarily follow that expected
from thermodynamic considerations. Recent studies
have shown that arsenite predominates in the oxidised
epilimnion of some stratified lakes whilst arsenate may
persist in the anoxic hypolimnion (Kuhn and Sigg, 1993;
Newman et al., 1998). Proportions of As species may also
vary according to the availability of particulate Fe and
Mn oxides (Pettine et al., 1992; Kuhn and Sigg, 1993).
Organic forms of As are usually minor in surface
waters. In lake waters from Ontario, Azcue and Nriagu
(1995) found As(III) concentrations of 7–75 mg l 1,
As(V) of 19–58 mg l 1 and only 0.01–1.5 mg l 1 of
organic As. Nonetheless, proportions of organic forms
of As can increase as a result of methylation reactions
catalysed by microbial activity (bacteria, yeasts, algae).
The dominant organic forms found are dimethylarsinic
acid
(DMAA;
(CH3)2AsO(OH))
and
monomethylarsonic acid (MMAA; CH3AsO(OH)2), where As
is present in both cases in the pentavalent oxidation
state. Proportions of these two species have been noted
to increase in summer as a result of increased microbial
activity (e.g. Hasegawa, 1997). The organic species may
also be more prevalent close to the sediment-water
interface (Hasegawa et al., 1999).
In groundwaters, the ratio of As(III) to As(V) can
vary greatly as a result of variations in the abundance of
redox-active solids, especially organic C, the activity of
microorganisms and the extent of convection and diffusion of O2 from the atmosphere. In strongly reducing
aquifers (Fe(III)- and SO4-reducing aquifers), As(III)
typically dominates, as expected from the redox sequence.
Reducing As-rich groundwaters from Bangladesh have
527
As(III)/AsT ratios varying between 0.1–0.9 but are typically around 0.5–0.6 (DPHE/BGS/MML, 1999; Smedley
et al., 2001b). Ratios in reducing groundwaters from Inner
Mongolia are typically 0.6–0.9 (Smedley et al., 2001a).
Concentrations of organic forms are generally low or
negligible in groundwaters (e.g. Chen et al., 1995).
2.4. Impact of redox kinetics on arsenic speciation
Redox reactions are important for controlling the
behaviour of many major and minor species in natural
waters, including that of As. However, in practice,
redox equilibrium is often achieved only slowly and the
redox potential tends to be controlled by the major elements (O, C, N, S and Fe). Redox-sensitive minor and
trace elements such as As respond to these changes
rather than control them. The slow rate of many heterogeneous redox reactions is supported by the studies
of Wersin et al. (1991) who estimated that the complete
reductive dissolution of Fe(III) oxides in an anoxic
Swiss lake sediment would take more than 1 ka. Equilibrium thermodynamic calculations predict that As(V)
concentrations should be greater than As(III) concentrations in all but strongly reducing conditions, i.e.
where SO4 reduction is occurring. While this is indeed
often found to be the case, such theoretical behaviour is
not necessarily followed quantitatively in natural waters
where different redox couples can point to different
implied redox potentials (Eh values), reflecting thermodynamic disequilibrium (Seyler and Martin, 1989; Eary
and Schramke, 1990; Kuhn and Sigg, 1993). In Oslofjord, Norway, As(III) was found under oxidising conditions (Abdullah et al., 1995). Also, in oxygenated
seawater, the As(V)/As(III) ratios should be of the order
of 1015–1026 (Andreae, 1979) whereas measured ratios
of 0.1–250 have been found, largely supported by biological transformations (Johnson and Pilson, 1975;
Cullen and Reimer, 1989). Oxidation of As(III) by dissolved O2, so-called oxygenation, is a particularly slow
reaction. For example, Johnson and Pilson (1975) gave
half-lives for the oxygenation of As(III) in seawater
ranging from several months to a year.
Other studies have demonstrated the stability of
As(V)/As(III) ratios over periods of days or weeks during water sampling when no particular care was taken to
prevent oxidation, again suggesting relatively slow oxidation rates. Andreae (1979) found stable ratios in seawater for up to 10 days (4 C). Cherry et al. (1979) found
from experimental studies that the As(V)/As(III) ratios
were stable in anoxic solutions for up to 3 weeks but
that gradual changes occurred over longer timescales.
Cherry et al. (1979) suggested that the measured As(V)/
As(III) ratios in natural waters, especially groundwaters,
might be used as an indicator of the ambient redox (Eh)
conditions as the redox changes are sufficiently rapid to
occur over periods of years. Yan et al. (2000) have also
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P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
concluded that the As(V)/As(III) ratio may be used as a
reliable redox indicator for groundwater systems. However, this optimism may be unfounded since Welch et al.
(1988) found that the Eh calculated from the As(V)–
As(III) couple neither agreed with that from the Fe(II)–
Fe(III) and other redox couples nor with the measured
Eh. Measurements of Eh in natural waters using Pt
electrodes are known to be problematic (Lindberg and
Runnells, 1984). The reliability of the As redox couple
as a redox indicator therefore remains to be seen. It is
clearly important that where such comparisons are
made, the Eh measurements are carried out without
disturbing the natural redox environment (Yan et al.,
2000). In cases where the aquifer is strongly stratified,
groundwater flow induced by pumping during sampling
or use may also lead to the mixing of waters with very
different redox potentials. Perhaps the most that can be
said at present is that the existence of As(III) implies
reducing conditions somewhere in the system.
Laboratory studies show that the kinetics of oxygenation of As(III) are slowest in the slightly acid range,
around pH 5 (Eary and Schramke, 1990) which is why
water samples are often acidified to about this pH to
preserve their in situ speciation. Eary and Schramke
(1990) also gave an empirical rate equation for the
reaction over the pH range 8–12.5. This was based on
the concentration (activity) of the H2AsO-3 species in
solution. They suggested that the half-life for As(III) in
natural waters is 1–3 a, although the rate may be greater
because of the presence of ‘unknown aqueous species’ or
oxide particles, especially Mn oxides. Certainly there is
considerable evidence that Mn oxides can increase the
rate of As(III) oxidation with half-lives being reduced to
as little as 10–20 min in the presence of Mn-oxide particles (Oscarson et al., 1981; Scott and Morgan, 1995).
This is used to advantage in the removal of As(III) from
drinking water (Driehaus et al., 1995). The rate of oxidation is independent of the concentration of dissolved O2
(Scott and Morgan, 1995), the rate being controlled by the
rate of a surface reaction. Less is known about the role of
Fe oxides in altering the oxygenation kinetics. Photochemical oxidation and reduction may be additional
factors in surface waters. Titanium-containing particles
may aid the photo-oxidation (Foster et al., 1998).
In the natural environment, the rates of both As(III)
oxidation and As(V) reduction reactions are controlled
by micro-organisms and can be orders of magnitude
greater than under abiotic conditions. For example,
sterile water samples have been observed to be less susceptible to speciation changes than non-sterile samples
(Cullen and Reimer, 1989). Wilkie and Hering (1998)
found that As(III) in geothermal waters input to streams
in SW USA oxidised rapidly downstream (pseudo firstorder half-life calculated at as little as 0.3 h) and they
attributed the fast rate to bacterial mediation. The
reduction of As(V) to As(III) in Mono Lake was also
rapidly being catalysed by bacteria with rate constants
ranging from 0.02 to 0.3 day 1 (Oremland et al., 2000).
Methylated As species are also readily oxidised chemically and biologically (Abdullah et al., 1995).
Less is known about the rate of solid-phase reduction
of As(V) to As(III) but there have been some studies
with soils and sediments. The evidence from soils is that
under moderately reducing conditions (Eh <100 mV)
induced by flooding, As(V) is reduced to As(III) in a
matter of days or weeks and adsorbed As(V) is released
as As(III) (Masscheleyn et al., 1991; Reynolds et al.,
1999). Masscheleyn et al. (1991) found from laboratory
experiments that some of the As was released before Fe,
implying reductive desorption from Fe oxides rather
than reductive dissolution. Up to 10% of the total As in
the soil eventually became soluble. Smith and Jaffé
(1998) modelled As(V) reduction in benthic sediments as
a first order reaction with respect to arsenate, with a rate
coefficient of 125 a 1.
3. Sources of arsenic
3.1. Minerals
3.1.1. Major arsenic minerals
Arsenic occurs as a major constituent in more than
200 minerals, including elemental As, arsenides, sulphides, oxides, arsenates and arsenites. A list of some of
the most common As minerals is given in Table 2. Most
are ore minerals or their alteration products. However,
these minerals are relatively rare in the natural environment. The greatest concentrations of these minerals
occur in mineralised areas and are found in close association with the transition metals as well as Cd, Pb, Ag,
Au, Sb, P, W and Mo. The most abundant As ore
mineral is arsenopyrite, FeAsS. It is generally believed
that arsenopyrite, together with the other dominant Assulphide minerals realgar and orpiment, are only formed
under high temperature conditions in the earth’s crust.
However, authigenic arsenopyrite has been reported in
sediments by Rittle et al. (1995) and orpiment has
recently been reported to have been formed by microbial precipitation (Newman et al., 1998). Although often
present in ore deposits, arsenopyrite is much less abundant than arsenian (‘As-rich’) pyrite (Fe(S,As)2) which
is probably the most important source of As in ore
zones (Nordstrom, 2000).
Where arsenopyrite is present in sulphide ores associated with sediment-hosted Au deposits, it tends to be
the earliest-formed mineral, derived from hydrothermal
solutions and formed at temperatures typically of 100 C
or more. This is followed by the formation of rarer
native As and thereafter arsenian pyrite. Realgar and
orpiment generally form later still. This paragenetic
sequence is often reflected by zonation within sulphide
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
529
Table 2
Major As minerals occurring in nature
Mineral
Composition
Occurrence
Native arsenic
Niccolite
Realgar
As
NiAs
AsS
Orpiment
Cobaltite
Arsenopyrite
Tennantite
Enargite
Arsenolite
As2S3
CoAsS
FeAsS
(Cu,Fe)12As4S13
Cu3AsS4
As2O3
Claudetite
Scorodite
Annabergite
Hoernesite
Haematolite
Conichalcite
Pharmacosiderite
Hydrothermal veins
Vein deposits and norites
Vein deposits, often associated with orpiment, clays and limestones, also deposits
from hot springs
Hydrothermal veins, hot springs, volcanic sublimation products
High-temperature deposits, metamorphic rocks
The most abundant As mineral, dominantly in mineral veins
Hydrothermal veins
Hydrothermal veins
Secondary mineral formed by oxidation of arsenopyrite, native arsenic and other
As minerals
Secondary mineral formed by oxidation of realgar, arsenopyrite and other As minerals
Secondary mineral
Secondary mineral
Secondary mineral, smelter wastes
As2O3
FeAsO4.2H2O
(Ni,Co)3(AsO4)2.8H2O
Mg3(AsO4)2.8H2O
(Mn,Mg)4Al(AsO4)(OH)8
CaCu(AsO4)(OH)
Secondary mineral
Fe3(AsO4)2(OH)3.5H2O
Oxidation product of arsenopyrite and other As minerals
minerals, with arsenopyrite cores zoning out to arsenian
pyrite and realgar-orpiment rims. Oxides and sulphates
are formed at the latest stages of ore mineralisation
(Arehart et al., 1993).
3.1.2. Rock-forming minerals
Though not a major component, As is also often
present in varying concentrations in other common
rock-forming minerals. As the chemistry of As follows
closely that of S, the greatest concentrations of the element tend to occur in sulphide minerals, of which pyrite
is the most abundant. Concentrations in pyrite, chalcopyrite, galena and marcasite can be very variable, even
within a given grain, but in some cases exceed 10 wt.%
(Table 3). Arsenic is present in the crystal structure of
many sulphide minerals as a substitute for S. Besides
being an important component of ore bodies, pyrite is
also formed in low-temperature sedimentary environments under reducing conditions. Such authigenic pyrite
plays a very important role in present-day geochemical
cycles. It is present in the sediments of many rivers,
lakes and the oceans as well as of many aquifers. Pyrite
commonly forms preferentially in zones of intense
reduction such as around buried plant roots or other
nuclei of decomposing organic matter. It is sometimes
present in a characteristic form as framboidal pyrite.
During the formation of this pyrite, it is likely that some
of the soluble As will also be incorporated. Pyrite is not
stable in aerobic systems and oxidises to Fe oxides with
the release of large amounts of SO4, acidity and associated trace constituents, including As. The presence of
pyrite as a minor constituent in sulfide-rich coals is ultimately responsible for the production of ‘acid rain’ and
acid mine drainage, and for the presence of As problems
around coal mines and areas of intensive coal burning.
High As concentrations are also found in many oxide
minerals and hydrous metal oxides, either as part of the
mineral structure or as sorbed species. Concentrations
in Fe oxides can also reach weight percent values
(Table 3), particularly where they form as the oxidation
products of primary Fe sulphide minerals, which have
an abundant supply of As. Adsorption of arsenate to
hydrous Fe oxides is particularly strong and sorbed
loadings can be appreciable even at very low As concentrations in solution (Goldberg, 1986; Manning and
Goldberg, 1996; Hiemstra and van Riemsdijk, 1996).
Adsorption to hydrous Al and Mn oxides may also be
important if these oxides are present in quantity (e.g.
Peterson and Carpenter, 1983; Brannon and Patrick,
1987). Arsenic may also be sorbed to the edges of clays
and on the surface of calcite (Goldberg and Glaubig,
1988), a common mineral in many sediments. However,
these loadings are much smaller on a weight basis than
for the Fe oxides. Adsorption reactions are responsible
for the relatively low (and non-toxic) concentrations of
As found in most natural waters.
Arsenic concentrations in phosphate minerals are
variable but can also reach high values, for example up to
1000 mg kg 1 in apatite (Table 3). However, phosphate
minerals are much less abundant than oxide minerals
and so make a correspondingly small contribution to
the As concentration in most sediments. Arsenic can
also substitute for Si4+, Al3+, Fe3+ and Ti4+ in many
mineral structures and is therefore present in many
other rock-forming minerals, albeit at much lower concentrations. Most common silicate minerals contain
530
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
Table 3
Typical As concentrations in common rock-forming minerals
Mineral
As concentration range (mg kg 1)
References
Sulphide minerals:
Pyrite
100–77,000
Pyrrhotite
Marcasite
Galena
Sphalerite
Chalcopyrite
5–100
20–126,000
5–10,000
5–17,000
10–5000
Baur and Onishi (1969);
Arehart et al. (1993); Fleet and Mumin (1997)
Boyle and Jonasson (1973);
Dudas (1984); Fleet and Mumin (1997)
Baur and Onishi (1969)
Baur and Onishi (1969)
Baur and Onishi (1969)
Oxide minerals
Haematite
Fe oxide (undifferentiated)
Fe(III) oxyhydroxide
Magnetite
Ilmenite
up to 160
up to 2000
up to 76,000
2.7–41
<1
Baur and Onishi (1969)
Boyle and Jonasson (1973)
Pichler et al. (1999)
Baur and Onishi (1969)
Baur and Onishi (1969)
Silicate minerals
Quartz
Feldspar
Biotite
Amphibole
Olivine
Pyroxene
0.4–1.3
<0.1–2.1
1.4
1.1–2.3
0.08–0.17
0.05–0.8
Baur and Onishi (1969)
Baur and Onishi (1969)
Baur and Onishi (1969)
Baur and Onishi (1969)
Baur and Onishi (1969)
Baur and Onishi (1969)
Carbonate minerals
Calcite
Dolomite
Siderite
1–8
<3
<3
Boyle and Jonasson (1973)
Boyle and Jonasson (1973)
Boyle and Jonasson (1973)
Sulphate minerals
Gypsum/anhydrite
Barite
Jarosite
<1–6
<1–12
34–1000
Boyle and Jonasson (1973)
Boyle and Jonasson (1973)
Boyle and Jonasson (1973)
Other minerals
Apatite
<1–1000
Halite
Fluorite
<3–30
<2
Baur and Onishi (1969),
Boyle and Jonasson (1973)
Stewart (1963)
Boyle and Jonasson (1973)
around 1 mg kg 1 or less. Carbonate minerals usually
contain less than 10 mg kg 1 As (Table 3).
3.2. Rocks, sediments and soils
3.2.1. Igneous rocks
Arsenic concentrations in igneous rocks are generally
low. Ure and Berrow (1982) quoted an average value of
1.5 mg kg 1 for all igneous rock types (undistinguished).
Averages for different types distinguished by silica content (Table 4) are slightly higher than this value but
generally less than 5 mg kg 1. Volcanic glasses are only
slightly higher with an average of around 5.9 mg kg 1
(Table 4). Overall, there is relatively little difference
between the different igneous rock types. Despite not
having exceptional concentrations of As, volcanic rocks,
especially ashes, are often implicated in the generation
of high-As waters (Nicolli et al., 1989; Smedley et al.,
2002). This may relate to the reactive nature of recent
acidic volcanic material, especially fine-grained ash and
its tendency to give rise to Na-rich high-pH groundwaters (Section 5).
3.2.2. Metamorphic rocks
Arsenic concentrations in metamorphic rocks tend to
reflect the concentrations in their igneous and sedimentary precursors. Most contain around 5 mg kg 1 or less.
Pelitic rocks (slates, phyllites) typically have the highest
concentrations with on average ca. 18 mg kg 1 (Table 4).
3.2.3. Sedimentary rocks
The concentration of As in sedimentary rocks is typically in the range 5–10 mg kg 1 (Webster, 1999), i.e.
slightly above average terrestrial abundance. Average
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
531
Table 4
Typical As concentrations in rocks, sediments, soils and other surficial deposits
Rock/sediment type
As concentration average and/
or range (mg kg 1)
No of
analyses
Igneous rocks
Ultrabasic rocks (peridotite, dunite,
kimberlite etc)
Basic rocks (basalt)
Basic rocks (gabbro, dolerite)
Intermediate (andesite, trachyte, latite)
Intermediate (diorite, granodiorite, syenite)
Acidic rocks (rhyolite)
Acidic rocks (granite, aplite)
Acidic rocks (pitchstone)
Volcanic glasses
1.5 (0.03–15.8)
40
2.3
1.5
2.7
1.0
4.3
1.3
1.7
5.9
78
112
30
39
2
116
Metamorphic rocks
Quartzite
Hornfels
Phyllite/slate
Schist/gneiss
Amphibolite and greenstone
5.5 (2.2–7.6)
5.5 (0.7–11)
18 (0.5–143)
1.1 (<0.1–18.5)
6.3 (0.4–45)
Sedimentary rocks
Marine shale/mudstone
Shale (Mid-Atlantic Ridge)
Non-marine shale/mudstone
Sandstone
Limestone/dolomite
Phosphorite
Iron formations and Fe-rich sediment
Evaporites (gypsum/anhydrite)
Coals
Bituminous shale (Kupferschiefer,
Germany)
Unconsolidated sediments
Various
Alluvial sand (Bangladesh)
Alluvial mud/clay (Bangladesh)
River bed sediments (Bangladesh)
Lake sediments, Lake Superior
Lake sediments, British Colombia
Glacial till, British Colombia
World average river sediments
Stream and lake silt (Canada)
Loess silts, Argentina
Continental margin sediments
(argillaceous, some anoxic)
Soils
Various
Peaty and bog soils
Acid sulphate soils (Vietnam)
Acid sulphate soils (Canada)
Soils near sulphide deposits
Contaminated surficial deposits
Mining-contaminated lake sediment,
British Colombia
(0.18–113)
(0.06–28)
(0.5–5.8)
(0.09–13.4)
(3.2–5.4)
(0.2–15)
(0.5–3.3)
(2.2–12.2)
3–15 (up to 490)
174 (48–361)
3.0–12
4.1 (0.6–120)
2.6 (0.1–20.1)
21 (0.4–188)
1–2900
3.5 (0.1–10)
0.3–35,000
100–900
3 (0.6–50)
2.9 (1.0–6.2)
6.5 (2.7–14.7)
1.2–5.9
2.0 (0.5–8.0)
5.5 (0.9–44)
9.2 (1.9–170)
5
6 (<1–72)
5.4–18
342 (80–1104)
Onishi and Sandell (1955);
Baur and Onishi (1969);
Boyle and Jonasson (1973);
Ure and Berrow (1982);
Riedel and Eikmann (1986)
12
4
2
75
16
45
15
40
205
45
5
13
23
119
310
2.3–8.2
7.2 (0.1–55)
13 (2–36)
6–41
1.5–45
126 (2–8000)
Reference
327
14
25
18
193
Boyle and Jonasson (1973)
Onishi and Sandell (1955);
Baur and Onishi (1969);
Boyle and Jonasson (1973);
Cronan (1972); Riedel and
Eikmann (1986); Welch et al.
(1988); Belkin et al. (2000)
Azcue and Nriagu (1995)
BGS and DPHE (2001)
BGS and DPHE (2001)
Datta and Subramanian (1997)
Allan and Ball (1990)
Cook et al. (1995)
Cook et al. (1995)
Martin and Whitfield (1983)
Boyle and Jonasson (1973)
Arribére et al. (1997);
Smedley et al. (2002)
Legeleux et al. (1994)
Boyle and Jonasson (1973)
Ure and Berrow (1982)
Gustafsson and Tin (1994)
Dudas (1984); Dudas et al. (1988)
Boyle and Jonasson (1973)
Azcue et al. (1994, 1995)
(continued on next page)
532
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
Table 4 (continued)
Rock/sediment type
As concentration average and/
or range (mg kg 1)
Mining-contaminated reservoir sediment, Montana
Mine tailings, British Colombia
Soils and tailings-contaminated soil, UK
Tailings-contaminated soil, Montana
Industrially polluted inter-tidal sediments, USA
Soils below chemicals factory, USA
100–800
903 (396–2000)
120–52,600
up to 1100
0.38–1260
1.3–4770
Sewage sludge
9.8 (2.4–39.6)
sediments are enriched in As relative to igneous rocks.
Sands and sandstones tend to have the lowest concentrations, reflecting the low As concentrations of their
dominant minerals: quartz and feldspars. Average sandstone As concentrations are around 4 mg kg 1 (Table 4)
although Ure and Berrow (1982) gave a lower average
value of 1 mg kg 1.
Argillaceous deposits have a broader range and
higher average As concentrations than sandstones, with
a typical average of around 13 mg kg 1 (Table 4; Ure and
Berrow, 1982). The higher values reflect the larger proportion of sulphide minerals, oxides, organic matter and
clays. Black shales have As concentrations at the high
end of the range, principally because of their enhanced
pyrite content. Data given in Table 4 suggest that marine argillaceous deposits have higher concentrations
than non-marine deposits. This may also be a reflection
of the grain-size distributions, with potential for a
higher proportion of fine material in offshore pelagic
sediments as well as systematic differences in sulphur
and pyrite contents. Marine shales tend to contain
higher S concentrations. Sediment provenance is also a
likely factor. Particularly high As concentrations have
been determined for shales from mid-ocean settings
(Mid-Atlantic Ridge average 174 mg kg 1; Table 4).
Concentrations in coals and bituminous deposits are
variable but often high. Samples of organic-rich shale
(Kupferschiefer) from Germany have As concentrations
of 100–900 mg kg 1 (Riedel and Eikmann, 1986;
Table 4). Some coal samples have been found with
extremely high concentrations up to 35,000 mg kg 1
(Belkin et al., 2000), although generally low concentrations of 2.5–17 mg kg 1 were reported by Palmer and
Klizas (1997). Carbonate rocks typically have low concentrations, reflecting the low concentrations of the
constituent minerals (ca. 3 mg kg 1; Table 4).
Some of the highest observed As concentrations are
found in ironstones and Fe-rich rocks. James (1966)
collated data for ironstones from various parts of the
world and reported As concentrations up to 800 mg
kg 1 in a chamosite-limonite oolite from the former
USSR. Boyle and Jonasson (1973) gave concentrations for
Fe-rich rocks up to 2900 mg kg 1 (Table 4). Phosphorites
No of
analyses
86
Reference
Moore et al. (1988)
Azcue et al. (1995)
Kavanagh et al. (1997)
Nagorski and Moore (1999)
Davis et al. (1997)
Hale et al. (1997)
Zhu and Tabatabai (1995)
are also relatively enriched in As, with values up to ca.
400 mg kg 1 having been reported.
3.2.4. Unconsolidated sediments
Concentrations of As in unconsolidated sediments are
not notably different from those in their indurated equivalents. Muds and clays usually have higher concentrations
than sands and carbonates. Values are typically 3–10 mg
kg 1, depending on texture and mineralogy (Table 4).
Elevated concentrations tend to reflect the amounts of
pyrite or Fe oxides present. Increases are also common
in mineralised areas. Placer deposits in streams can have
very high concentrations as a result of the abundance of
sulphide minerals. Average As concentrations for
stream sediments in England and Wales are in the range
5–8 mg kg 1 (AGRG, 1978). Similar concentrations
have also been found in river sediments where groundwater-As concentrations are high: Datta and Subramanian (1997) found concentrations in sediments
from the River Ganges averaging 2.0 mg kg 1 (range
1.2–2.6 mg kg 1), from the Brahmaputra River averaging 2.8 mg kg 1 (range 1.4–5.9 mg kg 1) and from the
Meghna River averaging 3.5 mg kg 1 (range 1.3–5.6 mg
kg 1).
Cook et al. (1995) found concentrations in lake sediments ranging between 0.9 and 44 mg kg 1 (median 5.5
mg kg 1) but noted that the highest concentrations were
present up to a few kilometres down-slope of mineralised areas. The upper baseline concentration for these
sediments is likely to be around 13 mg kg 1 (90th percentile). They also found concentrations in glacial till of
1.9–170 mg kg 1 (median 9.2 mg kg 1; Table 4) and
noted the highest concentrations down-ice of mineralised areas (upper baseline, 90th percentile, 22 mg
kg 1). Relative As enrichments have been observed in
reducing sediments in both nearshore and continentalshelf deposits (Peterson and Carpenter, 1986; Legeleux
et al., 1994). Legeleux et al. (1994) noted concentrations
increasing with depth (up to 30 cm) in continental shelf
sediments as a result of the generation of increasingly
reducing conditions. Concentrations varied between
sites, but generally increased with depth in the range
2.3–8.2 mg kg 1 (Table 4).
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
3.2.5. Soils
Baseline concentrations of As in soils are generally of
the order of 5–10 mg kg 1. Boyle and Jonasson (1973)
quoted an average baseline concentration in world soils
of 7.2 mg kg 1 (Table 4) and Shacklette et al. (1974)
quoted an average of 7.4 mg kg 1 (901 samples) for
American soils. Ure and Berrow (1982) gave a higher
average value of 11.3 mg kg 1. Peats and bog soils can
have higher concentrations (average 13 mg kg 1;
Table 4), principally because of increased prevalence of
sulphide mineral phases under the reduced conditions.
Acid sulphate soils which are generated by the oxidation
of pyrite in sulphide-rich terrains such as pyritic shales,
mineral veins and dewatered mangrove swamps can also
be relatively enriched in As. Dudas (1984) found As
concentrations up to 45 mg kg 1 in the B horizons of
acid sulphate soils derived from the weathering of pyrite-rich shales in Canada. Concentrations in the overlying leached (eluvial, E) horizons were low (1.5–8.0 mg
kg 1) as a result of volatilisation or leaching of As to
lower levels. Gustafsson and Tin (1994) found similarly
elevated concentrations (up to 41 mg kg 1) in acid sulphate soils from the Mekong delta of Vietnam.
Although the dominant source of As in soils is geological, and hence dependent to some extent on the
concentration in the parent rock material, additional
inputs may be derived locally from industrial sources
such as smelting and fossil-fuel combustion products
and agricultural sources such as pesticides and phosphate fertilisers. Ure and Berrow (1982) quoted concentrations in the range 366–732 mg kg 1 in orchard
soils as a result of the historical application of arsenical
pesticides to fruit crops.
3.2.6. Contaminated surficial deposits
Arsenic concentrations much higher than baseline
values have been found in sediments and soils contaminated by the products of mining activity, including
mine tailings and effluent. Concentrations in tailings piles
and tailings-contaminated soils can reach up to several
thousand mg kg 1 (Table 4). The high concentrations
reflect not only increased abundance of primary As-rich
sulphide minerals, but also secondary Fe arsenates and
Fe oxides formed as reaction products of the original ore
minerals. The primary sulphide minerals are susceptible
to oxidation in the tailings pile and the secondary
minerals have varying solubility in oxidising conditions
in groundwaters and surface waters. Scorodite (FeAsO4.2H2O) is a common sulphide oxidation product and
its solubility is considered to control As concentrations
in such oxidising sulphide environments. Scorodite is
metastable under most groundwater conditions and
tends to dissolve incongruently, forming Fe oxides and
releasing As into solution (Robins, 1987; Krause and
Ettel, 1989). In practice, a wide range of Fe–As mineral
solubility relationships are found which in part relate to
533
the mineral type (Krause and Ettel, 1989). There is some
confusion in the analysis of these solubility relationships
between congruent dissolution, incongruent dissolution
and sorption/desorption reactions. Secondary arsenolite
(As2O3) is also relatively soluble. Arsenic bound to Fe
oxides is relatively immobile, particularly under oxidising
conditions.
3.3. The atmosphere
The concentrations of As in the atmosphere are
usually low but as noted above, are increased by inputs
from smelting and other industrial operations, fossilfuel combustion and volcanic activity. Concentrations
amounting to around 10 5–10 3 mg m 3 have been
recorded in unpolluted areas, increasing to 0.003–0.18
mg m 3 in urban areas and greater than 1 mg m 3 close
to industrial plants (WHO, 2001). Much of the
atmospheric As is particulate. Total As deposition rates
have been calculated in the range < 1–1000 mg m 2 a 1
depending on the relative proportions of wet and dry
deposition and proximity to contamination sources
(Schroeder et al., 1987). Values in the range 38–266 mg
m 2 a 1 (29–55% as dry deposition) were estimated for
the mid-Atlantic coast (Scudlark and Church, 1988).
Airborne As is transferred to water bodies by wet or dry
deposition and may therefore increase the aqueous concentration slightly. However, there is little evidence to
suggest that atmospheric As poses a real health threat
for drinking-water sources. Atmospheric As arising
from coal burning has been invoked as a major cause of
lung cancer in parts of China (Guizhou Province), but
the threat is from direct inhalation of domestic coal-fire
smoke and especially from consumption of foods dried
over domestic coal fires, rather than from drinking water
affected by atmospheric inputs (Finkelman et al., 1999).
4. Mineral–water interactions
4.1. Controls on arsenic mobilisation
As with most trace metals, the concentration of As in
natural waters is probably normally controlled by some
form of solid-solution interaction. This is most clearly
the case for soil solutions, interstitial waters and
groundwaters where the solid/solution ratio is large but
it is also often true in open bodies of water (oceans,
lakes and reservoirs) where the concentration of solid
particles is small but still significant. In these open bodies,
the particles can be of mineral and biological origin. It is
likely that in most soils and aquifers, mineral–As interactions are likely to dominate over organic matter–As
interactions, although organic matter may interact to
some extent through its reactions with the surfaces of
minerals. Knowing the types of interaction involved is
534
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
important because this will govern the response of As to
changes in water chemistry. It will also determine the
modelling approach required for making predictions
about possible future changes and for understanding
past changes in As concentrations.
The importance of oxides in controlling the concentration of As in natural waters has been appreciated
for a long time (Livesey and Huang, 1981; Matisoff et al.,
1982) and there has been a wide range of studies to
measure the adsorption isotherms on natural and synthetic
oxide minerals and to establish the sorption processes at
the molecular scale (Table 5). Even so, important
uncertainties still remain in relation to the interactions
of As(III) and As(V) at environmental concentrations
and in the presence of other interacting ions. Frequently, the element which correlates best with As in
sediments is Fe. This is also the basis for the use of Fe
salts (as well as Al and Mn salts) in water treatment for
the removal of As and other elements (e.g. Edwards,
1994). The As content of residual sludges from water
treatment can be in the range 1000–10,000 mg kg 1
(Forstner and Haase, 1998; Driehaus et al., 1998). Clays
can also adsorb As(III) and As(V) (Manning and
Goldberg, 1997b) but their role in sediments in terms of
As binding is unclear at present.
It is difficult to study mineral-water interactions
directly in aquifers. Most studies, including those with a
bearing on As in groundwater, have been undertaken either
in soils, or in lake or ocean sediments and usually from
quite shallow depths. There is much to be learnt from the
studies of soils and sediments since the same general principles are expected to apply. One of the most important
areas where cross-fertilization of ideas can occur is in
understanding the behaviour of Fe oxides in reducing soils
and sediments and the influence of this on the release of As.
Matisoff et al. (1982) related reductive dissolution of Fe
oxides to the possible release of As in groundwater from an
alluvial aquifer in NE Ohio. Korte (1991) and Korte and
Fernando (1991) also speculated that desorption of As
from Fe oxides could occur in reducing, alluvial sediments
and that this could lead to high-As groundwaters.
mineral components, Fe oxides are probably the most
important adsorbents in sandy aquifers because of their
greater abundance and the strong binding affinity.
Nevertheless, Al oxides can also be expected to play a
significant role when present in quantity (Hingston et
al., 1971; Manning and Goldberg, 1997b). Experience
from water treatment suggests that below pH 7.5, Al
hydroxides are about as effective as Fe hydroxides (on a
molar basis) for adsorbing As(V) but that Fe salts are
more efficient at higher pH and for adsorbing As(III)
(Edwards, 1994).
The interactions of As with Fe oxides have been
studied in considerable detail in the laboratory and
therefore provide the best insight into the likely behavior of As-mineral interactions in aquifers. However,
most of these laboratory studies, particularly the older
studies, have been undertaken at rather high As concentrations (Hingston et al., 1971) and there is a paucity
of reliable adsorption data at the low mg l 1 level of
relevance to natural waters. In addition, there is uncertainty over the extent to which the Fe oxides most
commonly studied in the laboratory reflect the Fe oxides
found in the field. Field data for As(V) adsorption to
natural ‘diagenetic’ Fe oxides (captured in a lake with
vertically-installed Teflon sheets) closely paralleled the
laboratory data of Pierce and Moore (1982) which was
included in the Dzombak and Morel (1990) database (De
Vitre et al., 1991). However, it was considerably greater
than that calculated using Hingston et al.’s (1971) data
for As(V) adsorption on goethite, highlighting the high
affinity for As of freshly-formed ‘amorphous’ Fe oxides.
Paige et al. (1997) measured the As/Fe ratios during the
acid dissolution of a synthetic ferrihydrite containing
sorbed As(V) and concluded that the dissolution was
incongruent (i.e. Fe and As were not released in the
same proportion as found in the bulk mineral) and that
the initial As released was probably sorbed on the surface of the very small ferrihydrite particles. The same is
likely to happen during reductive dissolution. The
adsorbed As also slowed down the acid dissolution of
the ferrihydrite.
4.2. Arsenic associations in sediments
4.3. Reduced sediments and the role of iron oxides
The major minerals binding As (as both arsenate and
arsenite) in sediments are the metal oxides, particularly
those of Fe, Al and Mn (De Vitre et al., 1991; Sullivan
and Aller, 1996). About 50% of the Fe in freshwater
sediments is in the form of Fe oxides and about 20% of
the Fe is ‘reactive’ Fe. Clays also adsorb As because of
the oxide-like character of their edges. The extent of
As(V) sorption to, and coprecipitation on, carbonate
minerals is unknown but if it behaves like phosphate, it
is likely to be strongly retained by these minerals and
this may limit As concentrations in groundwaters from
limestone aquifers (Millero et al., 2001). Of these
A well-known sequence of reduction reactions occurs
when lakes, fjords, soils, sediments and aquifers become
anaerobic (Berner, 1981; Stumm and Morgan, 1995;
Langmuir, 1997). The processes causing changes in Fe
redox chemistry are particularly important since they
can directly affect the mobility of As. One of the principal causes of high As concentrations in subsurface
waters is the reductive dissolution of hydrous Fe oxides
and/or the release of adsorbed or combined As. This
sequence begins with the consumption of O2 and an
increase in dissolved CO2 from the decomposition of
organic matter. Next, NO-3 decreases by reduction to
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
535
Table 5
Studies of As adsorption by metal oxides
Mineral
Comment
Reference
Aluminium oxides
As(V) and As(III) adsorption on activated alumina: pH dependence,
kinetics, column breakthrough. Regeneration by desorbing with NaOH.
Modelling with pH-dependent Langmuir isotherm (for As) and surface
complexation model (for protons)
Ghosh and Yuan (1987)
‘Amorphous’
aluminium hydroxide
As(V) on precipitated Al(OH)3 (pH 3–10). ‘Adsorption’ exceeded
15 mol kg 1 at pH 5. Fitted data to pH dependent Langmuir isotherm
Anderson et al. (1976)
HFO
Kinetics and pH dependence of As(V) and As(III) adsorption on HFO
(202 m2 g 1). Found very high As(V) and As(III) loadings (up to 4–5 mol
As kg 1) at the highest concentrations. pH adsorption envelopes at
various AsT loadings
Raven et al. (1998)
HFO
Adsorption isotherms for arsenite and arsenate over free concentration
range from 10 7 to 10 3 M (pH 4–10). Fitted to Langmuir isotherm at
low concentrations and linear isotherm at higher concentrations.
Dzombak and Morel (1990) fitted this data to their diffuse double
layer model
Pierce and Moore (1982)
HFO
Sorption of As(V) and As(III) on HFO at As concentrations of
environmental significance (low micromolar range) and pH 4–9.
Compared results with Dzombak and Morel (1990) model predictions—
generally reasonable agreement. SO4 decreased adsorption of As(V) and
As(III), especially at low pH, while Ca increased As(V) adsorption at
high pH. 1 mM bicarbonate did not affect either As(V) or As(III)
adsorption greatly
Wilkie and Hering (1996)
HFO
A wide angle X-ray scattering (and EXAFS) study of two-line
ferrihydrite coprecipitated with varying amounts of As(V) suggested
that the As reduced crystallite size because of the formation of strongly
bound inner sphere complex between As(V) and edge sharing Fe(O,OH)6
octahedra. Saturation at As/Fe mol ratio of 0.68
Waychunas et al. (1996)
HFO
As(III) and As(V) adsorption and OH release/uptake on synthetic two-line
ferrihydrite. As(V) at pH 9.2 released up to 1 mol OH- per mol As sorbed
whereas As(III) released <0.25 mol As per mol Fe. At pH 4.6, OH release
was much less for As(V) adsorption and under these conditions there was a
net release of H+ by arsenite. These differences reflect the mechanism of
As adsorption and influence the pH dependence of adsorption
Jain et al. (1999)
Granular ‘ferric hydroxide’
(akageneite)
As(V) isotherms given in the sub-mM concentration range; SO4 competition
significant at mM concentrations below pH 7 only; phosphate competition
at ‘natural’ groundwater concentrations
Driehaus et al. (1998)
Goethite
An EXAFS and XANES study of As(III) adsorption to a synthetic goethite
suggested bidentate inner sphere binding. One plot of As(III) and As(V) pH
adsorption envelopes. As(III) data fitted to Constant Capacitance SCM
Manning et al. (1998)
Goethite
Batch adsorption of As(V) on synthetic goethite. Used Mo blue analysis
for As. Shows pH edge at about pH 9. Data fitted Langmuir isotherm
presumably at constant pH (up to 60 mg l 1 As)
Matis et al. (1997)
Goethite
Successfully applied the CD-MUSIC surface complexation model to
literature data for anion adsorption to goethite including As(V)-P
competition. The CD-MUSIC is the most promising of the SCMs for
modelling complex natural systems
Hiemstra and
van Riemsdijk (1999)
Goethite
As(V) adsorption on synthetic goethite primarily for a study of impact
on flocculation and electrokinetics. No isotherms. Final pH varied but
not defined
Matis et al. (1999)
(continued on next page)
536
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
Table 5 (continued)
Mineral
Comment
Reference
Goethite
EXAFS study of As(V) and Cr (VI) adsorption on goethite. Monodentate binding
favoured at low surface coverages of As(V), bidentate at high surface coverages
Fendorf et al. (1997)
Manganese oxides
As(III) and As(V) removal by MnO2(s) is similar, up to say 5 mmol As
mol 1 Mn at mM As equilibrium solution concentrations. Freundlich
isotherm obeyed. As(III) oxidised to As(V). Rapid oxidation (minutes)
and adsorption of As(III). Monitored Mn release and effect of pH, Ca,
phosphate and sulphate
Driehaus et al. (1995)
Birnessite, cryptomelane
and pyrolusite
Studied adsorption of As(III) and As(V) and kinetics of As(III) oxidation
in presence of various MnO2. As(III) adsorption (per unit weight of oxide):
cryptomelane> birnessite > pyrolusite whereas for As(V): cryptomelane
>pyrolusite >birnessite (not detectable). No isotherms given
Oscarson et al. (1983)
Goethite, hematite and
lepidocrocite
Batch adsorption of As(V), As(III), MMAA and DMAA on natural
minerals (coarse-grained and very low He-Ar surface area). As adsorption:
generally goethite >lepidocrocite > >hematite (pH 2–12, maximum often
pH 5–8). At pH 7 on goethite, As(III) >MMAA> DMAA >As(V) (?).
FA (up to 50 mg l 1) tended to reduce As adsorption. Gives Kd values
Bowell (1994)
Alumina, hematite,
quartz and kaolin
As(V) adsorption on natural, low surface area alumina, hematite, quartz
and kaolin (0.12–5 m2 g 1) at pH 3–10. Adsorption decreases with pH;
alumina=kaolin > hematite > > quartz. Gives Kd values and isotherms at
low concentrations. Some SO24 competition especially below pH 7. FA
( >10 mg l 1) generally reduced adsorption at pH 5–7 but not above pH 7
where FA is not adsorbed
Xu et al. (1988)
Alumina
On natural alumina, adsorption was As(V) > As(III) > MMAA=DMAA
(pH > 6). Maximum adsorption at pH 5 for As(V) and pH 7 for As(III).
As(V) but not As(III) adsorption decreased rapidly above pH 6. Log Kd
(l kg 1) at micromolar concentrations (pH 7) was 2.5–3.5 for As(V) and
about 1.5 for As(III). FA decreased adsorption
Xu et al. (1991)
Notes: HFO=hydrous ferric oxide.
SCM=surface complexation model.
EXAFS=extended X-ray absorption fine structure.
XANES=X-ray absorption near-edge structure.
MMAA=monomethylarsonic acid, CH3AsO(OH)2.
DMAA=dimethylarsinic acid, (CH3)2AsO(OH)).
FA=fulvic acid.
CD-MUSIC=charge distribution-multisite complexation model.
NO2 and the gases N2O and N2. Insoluble manganic
oxides dissolve by reduction to soluble Mn2+ and
hydrous ferric oxides are reduced to Fe2+. These processes are followed by SO24 reduction to S2 , then CH4
production from fermentation and methanogenesis, and
2
finally reduction of N2 to NH+
4 . During SO4 reduction,
2
the consequent S reacts with any available Fe to produce FeS and ultimately pyrite, FeS2. Iron is often more
abundant than S so that there is ‘excess Fe’ beyond that
which can be converted to pyrite (Widerlund and Ingri,
1995). Arsenic(V) reduction would normally be expected
to occur after Fe(III) reduction but before SO24 reduction.
In SO24 -poor environments, Fe from free Fe oxides is
solubilized as Fe2+ under reducing conditions. This gives
rise to characteristically high-Fe waters. Groundwaters
in such conditions tend to have Fe concentrations of 0.1–
30 mg l 1. The reaction is microbially mediated (Lovley
and Chappelle, 1995). There is also evidence for solid-state
transformations of the Fe oxides under reducing conditions. This is most obviously reflected in a colour change
from reddish/orange/brown/tan colours to grey/green/
blue colours. Changes to the magnetic properties have
also been documented (Sohlenius, 1996). Direct analysis
of the Fe(II) and Fe(III) contents of Fe oxides from
reduced lake waters and sediments often indicates the
presence of a mixed Fe(II)–Fe(III) oxide with an approximate average charge on the Fe of 2.5 (Davison, 1993).
The exact fate of Fe during reduction is not well understood, in part because it is probably very fine-grained
and difficult to observe directly. Mössbauer spectroscopy
is a useful technique for identifying the form of Fe oxides
in sediments, including anoxic sediments (Boughriet et
al., 1997; Drodt et al., 1997).
‘Green rusts’ are one possible product of the transformations. These have occasionally been identified or
suspected in anoxic soils and sediments (Taylor, 1980;
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
Boughriet et al., 1997; Cummings et al., 1999). They
consist of a range of green-coloured Fe(II)–Fe(III)
hydroxide minerals with a layered structure and a chargebalancing interlayer counterion, usually carbonate or sulphate. Green rusts were originally referred to as ‘hydrated
magnetite’ and given a composition ‘Fe3(OH)8’. Boughriet et al. (1997) suspected the presence of either greenrust-like compounds, Fe(III)–Fe(II)–(CO3)(OH) or
Fe(II)xCa1 xCO3 solid solutions, in anoxic sediments
from the Seine Estuary. They used 57Fe Mössbauer
spectroscopy to characterize the Fe. Green rusts have
also been identified in anaerobic soils and are thought to
play an important role in controlling soil solution Fe
concentrations (Genin et al., 1998).
Authigenic magnetite (Fe3O4), is another possible
product which has been identified in anaerobic sediments (Fredrickson et al., 1998), often with extremely
small particle sizes (Maher and Taylor, 1988; Canfield,
1989). Magnetite is frequently found in sediments as a
residual detrital phase from rock weathering but very
fine-grained magnetite is also formed by so-called ‘magnetotactic’ bacteria. Magnetite formation has been
established under reducing conditions in the laboratory
(Guerin and Blakemore, 1992). However, under strongly
reducing conditions magnetite is unstable and in the
presence of high concentrations of H2S, it slowly converts to pyrite over a period of 100a or more (Canfield
and Berner, 1987). At the sediment/water interface in
oceans, partial oxidation of primary magnetite (Fe3O4)
can lead to a coating of maghemite, g-Fe2O3. Further
burial and reduction leads to the dissolution of the primary
magnetite (Torii, 1997).
These studies of Fe oxides in reducing environments
indicate a lack of an as yet well-defined sequence of
events taking place when Fe(III) oxides are subjected to
strongly reducing conditions. The changes are evidently
substantial and can result in the partial dissolution of
the oxides and their transformation to completely new
mineral phases. It is not yet clear what impact these
transformations have on the adsorbed As load of the
original Fe(III) oxides. Suffice it to say that even quite
small changes in As binding could have a large impact
on porewater As concentrations because of the large
solid/solution ratio in sediments. Therefore, it is likely
that understanding the changes to the nature of Fe
oxide minerals in sedimentary environments is an
important aspect in understanding the processes leading
to As mobilisation in sedimentary environments.
4.4. Arsenic release from soils and sediments following
reduction
There is considerable evidence from laboratory studies that As is released from soils following flooding and
the development of anaerobic conditions (Deuel and
Swoboda, 1972; Hess and Blanchar, 1977; McGeehan
537
and Naylor, 1994; McGeehan, 1996; Reynolds et al.,
1999). Similar evidence is available from laboratory and
field studies of marine and lake sediments. Numerous
studies have demonstrated the release of both P (Mortimer, 1942; Farmer et al., 1994; Slomp et al., 1996) and As
(Aggett and O’Brien, 1985; Moore et al., 1988; De Vitre et
al., 1991; Azcue and Nriagu, 1995; Widerlund and Ingri,
1995) below the redox boundary in sediments. This release
has long been associated with Fe oxide dissolution.
Deuel and Swoboda (1972) found that reducing an
untreated black clay soil led to the release of As and
that the amount released was related to the total As
content of the soil and the redox potential. They proposed that the release was primarily due to reduction
(and dissolution) of ‘ferric arsenates’ rather than to
changes in the As speciation. Arsenic release occurred in
less than a week.
De Vitre et al. (1991) showed that there was a rapid
increase in porewater As concentrations (up to about 30
mg l 1) with depth in a lake sediment and that this was
mirrored by an increase in dissolved Fe. Upwardly diffusing Fe2+ was oxidised near the sediment-water
interface and precipitated as an Fe oxide which then
adsorbed the upwardly diffusing As. As noted in Section
2.2.7, much larger As concentrations, up to a peak of
6430 mg l 1, have been measured in porewaters from
shallow, anoxic lake sediments in the geothermal region
of New Zealand (Aggett and Kriegman, 1988). This
dissolved As, mostly As(III), diffused across the sediment-lake water interface and accumulated along with
dissolved Fe and Mn in the hypolimnion until turnover.
On the other hand, recent studies of As profiles in
Crowley Lake, California, a lake affected by large inputs
of As in surface water affected by geothermal sources,
suggest that As is not released from the sediments.
Rather it is taken up by the sediments, probably in solid
sulphide phases (Kneebone and Hering, 2000).
Guo et al. (1997) measured the rate of release of As
(and other trace elements) as a spiked sediment was
progressively reduced. Arsenic was rapidly released after
the Fe and Mn had dissolved, suggesting that dissolution rather than desorption was the dominant process,
or at least that dissolution and desorption occurred
simultaneously. Selective extractions suggested that
most of the As in the sediments was associated with Fe
and Mn oxides. Riedel et al. (1997) monitored the
release of metals when a column of estuarine sediment
was subjected to reducing conditions for several months.
Both As and Mn were released following reduction.
A few studies have attempted to differentiate between
the oxidation states of As sorbed by sediments. Masscheleyn et al. (1991) measured the release of As and
other metals following the flooding and reduction of an
As-contaminated soil and found that the release of some
As occurred before Fe dissolution but that the amount
of As released rapidly increased as the amount of Fe-
538
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
oxide dissolution increased. Both As(V) and As(III)
were released. Rochette et al. (1998) demonstrated with
XANES spectroscopy that reducing conditions can lead
to the conversion of As(V) to As(III) in the solid phase
of As minerals. Preliminary results based on XANES
also indicate a change in solid-state speciation of the As
in Bangladesh sediments in going from oxidising to
reducing conditions (Foster et al., 2000).
There is now considerable evidence that high-As
groundwaters can be associated with reducing conditions, particularly in alluvial and deltaic environments.
The Bengal Basin is the most notable example. While the
precise mechanisms responsible for this remain uncertain, it is possible that both reductive desorption and
reductive dissolution of As from oxides and clays play a
role. At the near-neutral pH of many groundwaters,
As(III) is expected to be less strongly sorbed than As(V)
and so some desorption of As may occur as a result of the
onset of strongly reducing conditions following sediment
burial (DPHE/BGS/MML, 1999, Vol. S4). Recent
experiments with synthetic Fe and Al oxides have
demonstrated the desorption of As following reduction
of As(V) to As(III) (Zobrist et al., 2000). Dissolution of
the oxides was not necessary for As release.
4.5. Speciation of elements in sediments and the role of
selective extraction techniques
The As concentration in sediments is normally too
low and/or the particles too small for direct investigation
of solid phase As speciation using techniques such as
XAFS and PIXE and so selective dissolution has been
most widely used. A number of chemical extraction
‘schemes’ have been devised which attempt to allocate
elements to particular solid phases but few of these are
specifically designed for speciating solid phase As.
Gómez-Ariza et al. (1998) developed a method to speciate
solid-phase As based on selective extraction of sediments with hydroxylamine hydrochloride, an acidic and
reducing extractant that is rather selective for extracting
Mn oxides but that also extracts small amounts of Fe
oxides. Hydroxylamine hydrochloride did not reduce
the As(V) during the extraction.
Keon et al. (2001) have described a more comprehensive 8-step sequential extraction scheme that attempts to
partition sediment As into various forms. The scheme
uses: (i) 1 M MgCl2 (for ionically-bound As); (ii) 1 M
NaH2PO4 (strongly adsorbed); (iii) 1 M HCl (acid
volatile sulphides, carbonates, Mn oxides and very
amorphous Fe oxyhydroxides); (iv) 0.2 M oxalate/oxalic
acid (amorphous Fe oxyhydroxides); (v) 0.05 M Ti(III)–
citrate-EDTA-bicarbonate (crystalline Fe oxyhydroxides); (vi) 10 M HF (arsenic oxides and silicates); (vii)
16 M HNO3 (pyrite and amorphous As2S3) and (viii) hot
16 M HNO3+30% H2O2 (orpiment and other recalcitrant minerals). About 90% of the As in a selection of
highly-contaminated sediments from the Wells G&H
Superfund site in the Aberjona catchment was found to
be in the (ii)–(iv) fractions described above.
Brannon and Patrick (1987) studied the kinetics of As
release and speciation [As(V), As(III), organic] from
freshwater sediments when incubated under both oxidising and reducing conditions. This included sediments
with and without added As(V). Most of the native and
added As was found in the ‘moderately reducible’ (oxalate-extractable) fraction. During incubation, a steady
release of As was observed over a 3-month period, with
As(V) occurring under oxidising conditions and As(III)
under reducing conditions. There was no concomitant
release of Fe (or Al or Mn), indicating that reductive
dissolution of Fe oxides was not responsible for the As
release. Brannon and Patrick (1987) speculated that a
change in the structure of the Fe oxides may have been
important. McGeehan (1996) was not sure whether the
As(V) reduction that occurs in flooded soils occurs in
the soil solution or on the soil particles.
Manning and Goldberg (1997a) measured As(V) and
As(III) adsorption by 3 Californian soils and found that
the soils with the highest citrate-dithionite-bicarbonate
extractable Fe and percent clay had the greatest affinity
for both As(III) and As(V). Arsenic(V) sorbed to a greater
extent than As(III) at the micromolar As concentrations
used, suggesting that As would be released under reducing
conditions when As(V) is reduced to As(III). Cummings
et al. (1999) found that As(V) was released from
hydrous ferric oxide (HFO) under reducing conditions
without any pre-reduction of As(V) to As(III). Scorodite, a ferric arsenate mineral (Table 2), was also in
part transformed to various ferrous arsenates. Therefore, the reduction of As(V) to As(III) may follow the
initial release of As into solution rather than initiate it.
Aggett and Roberts (1986) found a good correlation
between the release of both As and phosphate from
New Zealand lake sediments with that of EDTAextractable Fe. They concluded that the As was primarily associated with amorphous Fe oxides and they
suggested that because of the relative constancy of the
Fe/As ratio during dissolution, the As had probably
been coprecipitated with the Fe oxide rather than being
adsorbed to it.
Acid ammonium oxalate has been widely used for
extracting ‘amorphous Fe oxides’ and has been extensively used on sediments from Bangladesh (BGS and
DPHE, 2001). There was a good overall correlation
between extractable As and Fe but many other elements
were also correlated with the extractable Fe since this
tended to be most abundant in the fine-grained sediments. Nevertheless, the amount of oxalate-extractable
Fe and As of the aquifer sediments appeared to separate
low-As groundwater areas from high-As areas. Acid
ammonium oxalate is less selective with reducing sediments compared with oxidising sediments since mixed
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
valence Fe minerals such as magnetite are also dissolved
to a considerable extent. It also can dissolve appreciable
amounts of clays, as reflected by the extensive release of
Mg, Al and Si.
All of these studies demonstrate the ability of soils and
sediments to release As when subjected to reducing conditions but there is no clear consensus on the precise
mechanisms involved, particularly with respect to the
roles played by reductive desorption, reductive dissolution
and/or diagenetic changes to the mineral structure.
Alkaline extractants, particularly bicarbonate solutions at various pH values (e.g. 7.6, 8.5 and 10), have
been widely used to assess the plant availability of soil
phosphate. These have not yet been applied to the
extraction of As, specifically As(V), although Gustafsson and Jacks (1995) used 0.5 M NaOH. This
usually extracted some 1/2 to 2/3 of the ‘total’ As(V)
from the mineral horizons of some Swedish forest soils.
Unfortunately, none of the selective extraction
schemes is perfect or universally applicable and there is
little consensus on the best techniques to use. There are
no recognised procedures for solid phase As speciation,
although a number of traditional approaches have been
used on an ad hoc basis and new approaches are the
subject of current research (Keon et al., 2001). The
interpretation of results for many of these extractions is
difficult for minor and trace constituents which may be
released by dissolution, codissolution and desorption
processes. Nonetheless, these extractants can probe the
solid phase in a useful way that reflects to a varying extent
the nature of an element in the solid phase, and therefore
its potential behaviour or availability. In particular, such
techniques are useful for characterising very-fine grained
minerals or organic phases that are presently poorly
characterised by direct examination but which nevertheless play an important role in the behaviour of many
trace elements in the natural environment.
4.6. Transport of arsenic
The transport of chemicals and adsorption are closely
related in that adsorption slows down the transport of a
chemical compared with the water flow (Appelo and
Postma, 1994). In the simplest case of a linear adsorption
isotherm, this relationship is straightforward and the
partition coefficient, Kd, defines a constant retardation
factor. With a non-linear adsorption isotherm, the value
of Kd varies with concentration and is related to the
shape of the isotherm. Normally, the Kd decreases with
increasing concentration, leading to less retardation at
high concentrations and ultimately to self-sharpening
fronts and diffuse tails during transport. This means
that it is difficult to flush As completely from an aquifer.
Since the transport of a solute is closely related to the
extent of adsorption and the nature of the adsorption
isotherm, it follows that arsenate and arsenite, which
539
have different adsorption isotherms, should travel
through an aquifer with different velocities. This will
lead to their increased separation along a flow path.
This was demonstrated by Gulens et al. (1979) using
breakthrough experiments with columns of sand (containing 0.6% Fe and 0.01% Mn) and various groundwaters pumped continuously from piezometers. They
studied As(III) and As(V) mobility with groundwaters
having a range of Eh and pH values using radioactive
74
As (half-life=17.7 days) and 76As (half-life=26.4 h)
to monitor the breakthrough of As. They showed that:
(i) As(III) moved 5–6 times faster than As(V) under
oxidising conditions (pH 5.7); (ii) with a ‘neutral’
groundwater (pH 6.9), As(V) moved much faster than
under condition (i) but was still slower than As(III); (iii)
with reducing groundwater at pH 8.3, both As(III) and
As(V) moved rapidly through the column; (iv) when the
amount of As injected was substantially reduced, the
mobility of the As(III) and As(V) was greatly reduced.
This chromatographic effect may account in part for the
highly variable As(III)/As(V) ratios found in many
reducing aquifers. Such a separation is used to advantage in analytical chemistry to speciate As with various
columns. Chromatographic separation during transport
will also tend to uncorrelate any correlations found at
the source, for example in the As versus Fe relationship,
thus further complicating a simple interpretation of well
water analyses.
A number of experimental investigations of As
adsorption reported in the literature, most commonly
for pure Fe and Al oxides, allow Kd values to be estimated for a range of adsorbents (Table 6). The Kds thus
derived, here mostly calculated for pH 7, have a large
range, spanning around 6 orders of magnitude. All other
things being equal, Kds tend to decrease with increasing
As concentration reflecting some nonlinearity in the
sorption isotherm. As expected, under similar conditions, Kds are often greater for As(V) than for As(III)
but this is not always the case. Values are highest for
HFO which in part reflects the large adsorption capacity
of HFO. High Kd values are also found for Mn oxides
(birnessite). Experiments with Mn oxides by Driehaus et
al. (1995) suggest that values are high for both As(V)
and As(III), although in practice, much of the As(III)
was probably oxidised to As(V) during the experiment
by the Mn(IV) oxide (Oscarson et al., 1981; 1983). There
is a need for more high-quality data on the sorption of
As(V) and As(III) at low As concentrations (1–100 mg
l 1) including experiments carried out in the presence of
competing anions and using aquifer solids rather than
synthetic minerals.
A few field-based investigations have been carried out
on natural and contaminated systems which allow Kd
values for As sorption to be determined directly. Kuhlmeier (1997a,b) studied the transport of As in highlycontaminated clayey and sandy soils from around an
540
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
Table 6
Solid-solution partition coefficients (Kds) calculated from experimental data at or near pH 7 for a range of oxides and clays
Material
Components
Kd (l kg 1)
pH
Comments
Reference
Alumina (a-Al2O3)
As(III)
42
7
Xu et al. (1991), Fig. 2
Alumina (a-Al2O3)
As(V)
760
7
Alumina (a-Al2O3)
As(V)
520
7
Amorphous aluminium
hydroxide
Gibbsite
Gibbsite
As(III)
230
7
a
c=37 mg l 1,
I=0.1 M
c=3.8 mg l 1,
I=0.1 M
c=1200 mg l 1,
I=0.01 M
c=4 mg l 1
As(V)
As(V)
133
32
7
7
Hingston et al. (1971), Fig. 3a
Hingston et al. (1971), Fig. 3c+b
Goethite
As(V)
192
7
Goethite
As(V)+P
54
7
Goethite (natural)
Goethite (natural)
Haematite
As(III)
As(V)
As(V)
32
1800
34
7
7
7
Haematite (natural)
Haematite (natural)
HFO
HFO
As(III)
As(V)
As
As(III)
21
25
7000
120,000
7
7
HFO
As(III)
670,000
7
HFO
As(III)
7340
7
HFO
As(III)
520,000
7
HFO
As(III)+Si
13,000
7
HFO
As(V)
>1,000,000
7
HFO
As(V)
460,000
7
HFO
As(V)
120,000
7
HFO
As(V)
37,000
7
HFO
As(V)
66,000
7
HFO
As(V)+Si
8,100
7
HFO (granular)
Iron-oxide-coated sand
As(V)
Oxidn state
uncertain
As(III)
As(V)
2,100,000
600
6
?
c=16 mg l 1
c=16 mg/l+
5.5 mg l 1 P
c=4.9 mg l 1,
I=0.1 M
c=66 mg l 1,
+74 mg l 1 P,
I=0.1 M
c=40 mg l 1
c=1.5 mg l 1
c=48 mg l 1,
I=0.1 M
c=48 mg l 1
c=45 mg l 1
c=50 mg l 1
c=310 mg l 1,
I=0.1 M
c=27 mg l 1,
I=0.01 M
c=18900 mg l 1,
I=0.01 M
c=30 mg l 1,
I=0.01 M; possibly
partial oxidation
during experiment
c=1690 mg l 1,
SiT=62 mg l 1,
I=0.1 M
c <50 mg l 1,
I=0.1 M
c=32 mg l 1
I=0.01 M
c=160 mg l 1
I=0.01 M
c=850 mg l 1
I=0.01 M
c=19500 mg l 1
I=0.01 M
c=2130 mg l 1,
SiT=62 mg l 1,
I=0.1 M
50 mg l 1
c=50 mg l 1
35
1000
7
7
c=39 mg l
c=2.7 mg l
Lepidocrocite (natural)
Lepidocrocite (natural)
7
1
1
Xu et al. (1988), Fig. 3
Anderson et al. (1976), Fig. 2
Manning and Goldberg (1997a), Fig. 1
Hingston et al. (1971), Fig. 2
Hingston et al. (1971), Fig. 4
Bowell (1994), Fig. 2
Bowell (1994), Fig. 2
Xu et al. (1988), Fig. 3
Bowell (1994), Fig. 2
Bowell (1994), Fig. 2
Thirunavukukkarasu et al. (2001), Fig. 3
Swedlund and Webster (1999), Fig. 1a
Pierce and Moore (1982), Fig. 1
Pierce and Moore (1982), Fig. 3
Wilkie and Hering (1996), Fig. 1
Swedlund and Webster (1998), Fig. 1a
Swedlund and Webster (1998), Fig. 1b
Pierce and Moore (1982), Fig. 5
Pierce and Moore (1982), Fig. 5
Pierce and Moore (1982), Fig. 5
Pierce and Moore (1982), Fig. 7
Swedlund and Webster (1998), Fig. 1b
Driehaus et al. (1998), Fig. 1
Thirunavukukkarasu et al. (2001), Fig. 2
Bowell (1994), Fig. 2
Bowell (1994), Fig. 2
(continued on next page)
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
541
Table 6 (continued)
Material
Components
Kd (l kg 1)
pH
Comments
Reference
Birnessite (d-MnO2)
‘‘As(III)’’
46,000
7
Driehaus et al. (1995), Fig. 1
Birnessite (d-MnO2)
As(V)
57,500
7
Quartz
As(V)
2
7
Illite
Kaolinite
Kaolinite
As(III)
As(III)
As(V)
98
19
760
7
7
7
Wyoming Bentonite
As(III)
30
7
c=75 mg l 1,
I=0.01 M; As(III)
probably all oxidised
during experiment
c=75 mg l 1,
I=0.01 M
c=71 mg l 1,
I=0.1 M
c=9 mg l 1
c=20 mg l 1
c=3.8 mg l 1,
I=0.1 M
c=17 mg l 1
a
Driehaus et al. (1995), Fig. 1
Xu et al. (1988), Fig. 3
Manning and Goldberg (1997a), Fig. 1
Manning and Goldberg (1997a), Fig. 1
Xu et al. (1988), Fig. 3
Manning and Goldberg (1997a), Fig. 1
c=final dissolved As concentration; I=ionic strength.
old As herbicide plant in Houston, Texas. He used column experiments to estimate ‘apparent’ Kd values.
These were time- and implicitly concentration-dependent and for the sandy soils ranged from 0.26 l kg 1
after one void volume to 3.3 l kg 1 after 6 void volumes.
They were not too different for the clayey materials.
However, the overall As concentrations were very high:
the groundwater was heavily contaminated with As
(408–464 mg l 1), mostly as MMAA. The sediment
contained only a few mg kg 1 of inorganic As.
Baes and Sharp (1983) gave Kd values of 1.0–8.3 l
kg 1 (median 3.3) for As(III) binding by soils and 1.9–
18.0 l kg 1 (median 6.7) for As(V). The authors have
estimated Kds for 3 Californian soils (pH 5.7–7.1) at an
equilibrium concentration of 50 mg l 1 to range from 5–
52 l kg 1 for As(III) and 10–80 for As(V) (Manning and
Goldberg, 1997a, Fig. 4).
Data from Sullivan and Aller (1996) indicate that Kd
values calculated for sediment profiles from the Amazon
Shelf are in the approximate range 11–5000 l kg 1.
High-As porewaters were mostly from the zones with
low Kd values (typically < 100 l kg 1). The concentrations of reactive solid As derived from an acid leach
were in the approximate range 1–13 mg kg 1 and the
authors gave an average total sediment As concentration of 12 mg kg 1. Smedley et al. (2000) also estimated
Kd for loess sediments bearing high-As groundwaters
from La Pampa, Argentina. They estimated a low value,
close to 1 l kg 1, using coupled porewater and oxalateextract data. Arsenic concentrations of the analysed
sediments were in the range 3–18 mg kg 1.
Calculated sorption and retardation factors, based on
the Dzombak and Morel (1990) diffuse double layer
model and database for HFO, indicate highly variable
Kds depending on: the HFO (and other active oxide)
content of the sediment, As speciation, As concentra-
tion, pH and concentration of competitors such as
phosphate (DPHE/BGS/MML, 1999, Vol. S3 and S4).
Inferred Kds from these calculations for Bangladesh
sediments range from about 1 l kg 1 for As(V) under
‘high P-low Fe’ conditions to more than 200 l kg 1
under ‘low P-high Fe’ conditions. Broadly similar values
are inferred for As(III) sorption under comparable conditions—the presence of high phosphate concentrations
is calculated to severely reduce As(V) sorption but not
As(III) sorption. Coupled studies of sediment and
groundwater chemistry in piezometers (10–50 m depth)
in 3 high-As groundwater locations in Bangladesh (BGS
and DPHE, 2001) suggest that the Kds for As are low,
approximately 2–6 l kg 1. These values have been
derived by assuming that the sorbed As can be equated
to the oxalate-extractable As and so are likely to represent an upper estimate.
Values obtained for Kd from these field studies in
areas having high-As groundwaters generally suggest
that it is, for whatever reason, the low Kd values that are
responsible for the high dissolved As concentrations,
rather than high absolute concentrations of As in the
sediment. The Kd values given in Table 6 are in many
cases much higher than found in natural sediments, even
after allowing for the relatively minor concentration (a
few percent) of Fe oxides in most sediments. There is some
evidence that use of the surface complexation model and
database of Dzombak and Morel (1990) can lead to an
overestimation of sorption when applied to natural sediments (Kent et al., 1995). This may reflect a difference
between the natural Fe oxides found in sediments and
synthetic HFO used for most laboratory studies or other
factors such as the presence of competing ions in the
natural sediments. Certainly it would be difficult to infer
the location of high-As groundwaters from an analysis
of total sediment As concentrations alone.
542
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
5. Groundwater environments with high arsenic
concentrations
5.1. World distribution of groundwater arsenic problems
A number of large aquifers in various parts of the
world have been identified with problems from As occurring at concentrations above 50 mg l 1, often significantly
so. The most noteworthy occurrences are in parts of
Argentina, Bangladesh, Chile, China, Hungary, India
(West Bengal), Mexico, Romania, Taiwan, Vietnam and
many parts of the USA, particularly the SW (Fig. 3).
Some of the better documented cases are summarised in
Table 7. These include natural sources of enrichment as
well as mining-related sources. Recent reconnaissance
surveys of groundwater quality in other areas such as parts
of Nepal, Myanmar and Cambodia have also revealed
concentrations of As in some sources exceeding 50 mg l 1,
although documentation of the affected aquifers is so far
limited. Arsenic associated with geothermal waters has
also been reported in several areas, including hot springs
from parts of Argentina, Japan, New Zealand, Chile,
Kamchatka, Iceland, France, Dominica and the USA.
Localised groundwater As problems are now being
reported from an increasing number of countries and
many new cases are likely to be discovered. Until
recently, As was not traditionally on the list of elements
routinely tested by water-quality testing laboratories
and so many high-As water sources may have been
missed. Revision of drinking-water regulations and
guidelines for As has prompted a reassessment of the
situation in many countries. The recent discovery of As
enrichment on a large scale in Bangladesh has highlighted the need for a rapid assessment of the situation
in alluvial aquifers worldwide. As described above, As
problems also occur in some areas where sulphidemining activity is prevalent, the As being released from
sulphide minerals as they are oxidised as a result of
mining operations. In mining areas, As problems can be
severe with concentrations in affected waters sometimes
being in the mg l 1 range. However, unlike As occurrences
in major aquifers, the problems in these areas are typically
localised, rather than of widespread occurrence. Miningrelated As problems in water have been identified in
many parts of the world, including Ghana, Greece,
Thailand and the USA (Fig. 3).
While high-As groundwaters (with As above drinking-water standards) are not uncommon, they are by no
means typical of most aquifers and only exist under special circumstances. These relate to both the geochemical
environment and to the past and present hydrogeology.
Paradoxically, high-As groundwaters are not necessarily
related to areas of high As concentrations in the source
rocks. Distinctive groundwater As problems occur
under both reducing and oxidising groundwater conditions; also under both humid/temperate and arid climates. Below, the authors discuss the characteristics of
the As problems worldwide through a series of type
examples. These have been ordered according to the
environment under which they are developed.
Fig. 3. Distribution of documented world problems with As in groundwater in major aquifers as well as water and environmental
problems related to mining and geothermal sources. Areas in blue are lakes.
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
543
Fig. 4. Smoothed map showing the distribution of As in groundwater from shallow ( <150 m) tubewells in Bangladesh.
5.2. Reducing environments
5.2.1. Bangladesh and India (West Bengal)
In terms of the population exposed, As problems in
groundwater from the alluvial and deltaic aquifers of
Bangladesh and West Bengal represent the most serious
occurrences identified globally. Concentrations in
groundwaters from the affected areas have a very large
range from < 0.5 to ca. 3200 mg l 1 (DPHE/BGS/MML,
1999; CGWB, 1999; BGS and DPHE, 2001). Some 27%
of shallow (< 150 m deep) wells in Bangladesh contain
more than 50 mg l 1 As (BGS and DPHE, 2001). The
worst-affected area is in the SE of Bangladesh (Fig. 4)
where in some districts more than 90% of wells are
affected. Resultant health problems were first identified
in West Bengal in the 1980s but the first diagnosis in
Bangladesh was not made until 1993. Up to around 30–
35 million people in Bangladesh are estimated to be
exposed to As in drinking water at concentrations above
50 mg l 1 (BGS and DPHE, 2001) and up to 6 million in
West Bengal (Table 7). Skin disorders including hyper/
hypopigmentation changes and keratosis are the most
common external manifestations, although skin cancer
has also been identified. Around 5000 patients have
been identified with As-related health problems in West
Bengal (including skin pigmentation changes) although
some estimates put the number of patients with arsenicosis at more than 200,000 (Smith et al., 2000). The
number in Bangladesh is not known but must be many
times greater than in West Bengal. The instance of
internal As-related health problems is also not known
but could be appreciable.
West Bengal and Bangladesh rely heavily on groundwater for public drinking-water supply. Groundwater
development has been actively encouraged in the region
over the last few decades by government and other
agencies as a means of providing an alternative to polluted surface water and thereby reducing the incidence
of water-borne diseases. There has also been a rapid
increase in the number of private tubewells and now the
number of these exceeds the number of public tubewells.
In this sense, the increase in use of groundwater has
been very successful. The identification of chronic health
problems related to As was unforeseen and has taken a
number of years to become apparent.
The affected aquifers are generally shallow (less than
100–150 m deep), of Holocene age and comprise micaceous sands, silts and clays deposited by the Ganges,
Brahmaputra and Meghna river systems and their precursors. The sediments are derived from the upland
Himalayan catchments and from basement complexes
of the northern and western parts of West Bengal. In
most affected areas, the aquifer sediments are capped by
a layer of clay or silt (of variable thickness) which
effectively restricts entry of air to the aquifers. This,
together with the presence of recent solid organic matter
deposited with the sediments, has resulted in the development of highly reducing conditions which favour the
mobilisation of As. The mobilisation has probably
occurred by a complex combination of redox changes
544
Table 7
Summary of documented cases of naturally-occurring As problems in world groundwaters (includes some mining cases)
Area (km2)
Population Concentration Aquifer type
exposeda
ranges (mg l 1)
Groundwater conditions
Reference
Bangladesh
150,000
ca. 3107
<0.5–2500
23,000
6106
<10–3200
Strongly reducing, neutral pH,
high alkalinity, slow groundwater
flow rates
As Bangladesh
DPHE/BGS/MML (1999);
BGS and DPHE (2001)
West Bengal
Holocene alluvial/deltaic
sediments. Abundance of
solid organic matter
As Bangladesh
China:
Taiwan
4000
5.6106
?105
(formerly)
10–1820
Sediments, including
black shale
4300 (HB)
? ca. 105
?30,000 total in HB
<1–2400
Holocene alluvial and
lacustrine sediments
Strongly reducing, artesian
conditions, some groundwaters
contain humic acid
Strongly reducing conditions,
neutral pH, high alkalinity.
Deep groundwaters often artesian,
some have high concentrations
of humic acid
38,000
? 500
diagnosed
40–750
Holocene alluvial plain
1200
identified
110,000
>106
1–3050
29,000
<2–176
106
2106
<1–5300
(7500 in some
porewaters)
Holocene and earlier loess
with rhyolitic volcanic ash
125,000
500,000
100–1000
?Quaternary volcanogenic
sediment
Inner Mongolia
(Tumet Plain
including Huhhot
Basin (HB); Ba
Men, Bayingao,
Hexi)
Xinjiang (Tianshan
Plain)
Shanxi
Red River delta,
Vietnam
Hungary, Romania
(Danube Basin)
Argentina (ChacoPampean Plain)
Northern Chile
(Antofagasta)
3.5x105
(total)
South-west USA:
Basin and Range,
Arizona
Alluvial plain
Holocene alluvial/deltaic
sediments
Quaternary alluvial plain
200,000
Reducing, deep wells (up to 660 m)
are artesian
?Reducing
Reducing, high Fe, Mn, NH4,
high alkalinity
Reducing groundwater, some
artesian. Some high in humic acid
Oxidising, neutral to high pH,
high alkalinity. Groundwaters
often saline. As mainly present as
As(V), accompanied by high B,
V, Mo, U. Also high As
concentrations in some river waters
Generally oxidising. Arid
conditions, high salinity, high B.
Also high-As river waters
CGWB (1999); PHED/
UNICEF (1999)
Sun et al. (2001)
Kuo (1968), Tseng et al. (1968)
Luo et al. (1997), Zhai et al.
(1998), Ma et al. (1999), Sun et al.
(1999), Smedley et al. (2001a)
Wang and Huang (1994)
Sun et al. (1999)
Berg et al. (2001)
Varsányi et al. (1991), Gurzau and
Gurzau (2001)
Nicolli et al. (1989), Sancha and
Castro (2001), Smedley et al. (2002)
Cáceres et al. (1992), Karcher et al.
(1999); Sancha and Castro (2001)
Smith et al. (1992)
up to 1300
Alluvial basins, some
evaporites
Oxidising, high pH. As
Robertson (1989)
(mainly As(V)) correlates positively
with Mo, Se, V, F
(continued on next page)
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
Country/region
Table 7 (continued)
Country/region
Area (km2)
Population Concentration Aquifer type
exposeda
ranges (mg l 1)
<1–2600
Holocene and older
basin-fill sediments
Southern Carson
Desert, Nevada
up to 2600
Holocene mixed aeolian,
alluvial, lacustrine sediments,
some thin volcanic ash bands
1300
Salton Sea Basin
Mexico (Lagunera)
32,000
4105
8–620
Volcanic sediments
Reference
Internally-drained basin. Mixed
redox conditions Proportion of
As(III) increases with well depth.
High salinity in some shallow
groundwaters. High. Se, U, B, Mo
Largely reducing, some high pH.
Some with high salinity due to
evaporation. Associated high U,
P, Mn, DOC (Fe to a lesser extent)
Some saline groundwaters, with
high U
Oxidising, neutral to high pH,
As mainly as As(V)
Fujii and Swain (1995)
Some problem areas related to mining activity and mineralised areas
Thailand
100
?15,000
1 to 5000
Dredged Quaternary alluvium
Oxidation of disseminated
(Ron Phibun)
(some problems in limestone), tailings arsenopyrite due to former tin
mining, subsequent groundwater
rebound
Greece (Lavrion)
Mine tailings
Mining
Fairbanks,
up to 10,000
Schist, alluvium, mine tailings
Gold mining, arsenopyite,
Alaska, USA
possibly scorodite
Moira Lake,
100
50–3000
Mine tailings
Ore mining (Au, haematite,
Ontario, Canada
magnetite, Pb, Co)
Coeur d’Alene,
up to 1400
Valley-fill deposits
River water and groundwater
Idaho, USA
affected by Pb-Zn-Ag mining
Lake Oahe,
up to 2000
Lake sediments
Arsenic in sediment porewaters
South Dakota, USA
from Au mining in the Black Hills
Bowen Island,
50
0.5–580
Sulphide mineral veins in
Neutral to high-pH groundwaters
British Colombia
volcanic country rocks
(up to 8.9), As correlated with B, F
Northern Bavaria,
2500
<10–150
Upper Triassic Keuper Sandstone
Variable SO4 concentrations,
Germany
with mineralisation (uncharacterized) usually low NO3, low dissolved
O2 concentrations
a
Exposed refers to population drinking water with As >50 mg l
1
Welch and Lico (1998)
Welch and Lico (1998)
Del Razo et al. (1990)
Williams et al. (1996),
Williams (1997)
Wilson and Hawkins (1978);
Welch et al. (1988)
Azcue and Nriagu (1995)
Welch et al. (1988), Mok and
Wai (1990)
Ficklin and Callender (1989)
Boyle et al. (1998)
Heinrichs and Udluft (1999)
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
Tulare Basin,
5000
San Joaquin Valley,
California
Groundwater conditions
(drinking-water standard of most countries).
545
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P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
brought on by rapid burial of the alluvial and deltaic
sediments, including reduction of the solid-phase As to
As(III), desorption of As from Fe oxides, reductive dissolution of the oxides themselves and likely changes in
Fe-oxide structure and surface properties following the
onset of reducing conditions (BGS and DPHE, 2001).
Deep wells, tapping depths greater than 150–200 m,
almost invariably have low As concentrations: less than
5 mg l 1 and usually less than 0.5 mg l 1 (BGS and
DPHE, 2001). Also wells from the older Plio-Pleistocene
sediments of the Barind and Madhupur Tracts have low
As concentrations (Fig. 4). It is a fortunate fact that
both Calcutta and Dhaka draw their water from older
sediments and do not have an As problem. Dhaka is
sited at the southern tip of the Madhupur Tract (Fig. 4).
Dug wells also generally have low As concentrations,
typically < 10 mg l 1 (BGS and DPHE, 2001).
The characteristic chemical features of the high-As
groundwaters of the Bengal Basin are high Fe ( > 0.2 mg
l 1), Mn ( > 0.5 mg l 1), HCO3 ( > 500 mg l 1) and often
P (> 0.5 mg l 1) concentrations, and low Cl ( < 60 mg
l 1), SO24 (< 1 mg l 1), NO3 and F ( < 1 mg l 1) concentrations, with pH values close to or greater than 7.
Measured redox potentials are typically less than 100
mV (PHED, 1991; CGWB, 1999; DPHE/BGS/MML,
1999; BGS and DPHE, 2001). However, the correlations
between dissolved elements are usually far from perfect
and where good correlations with As are found, these
are only applicable locally and are of limited value for
quantitative prediction of As concentrations, even at a
village scale. For example, some workers have found a
positive correlation between As and Fe in village studies
(e.g. Nag et al., 1996), but this is not true of Bangladesh
and West Bengal as a whole. One commonly observed
relationship in the groundwaters is a general negative
correlation between As and SO4 concentrations (BGS
and DPHE, 2001). This association suggests that As
mobilisation is effected under the most strongly reducing conditions, coincident with SO24 reduction. Some
of the groundwaters of Bangladesh are sufficiently
reducing for CH4 generation to have taken place.
Arsenic speciation studies have revealed a large range
in the relative proportions of dissolved arsenate and
arsenite present in the groundwaters (e.g. Das et al.,
1995; Acharyya, 1997). The modal proportion of
arsenite appears to be between 50 and 60% of the total
As (BGS and DPHE, 2001). This may reflect lack of
redox equilibrium in the aquifer or a mixed groundwater from a strongly stratified aquifer. Low As concentrations are most common in groundwaters from
northern Bangladesh and the aquifers in Plio-Pleistocene sediments of the uplifted Barind and Madhupur
Tracts of north-central Bangladesh.
The regional distribution of the high-As waters in
West Bengal and Bangladesh is known to be extremely
patchy (PHED, 1991; CSME, 1997; BGS and DPHE,
2001), presumably in part because of great variation in
sedimentary characteristics and variations in abstraction
depth. Estimates of the proportions of tubewells affected in West Bengal are not well-documented and difficult to assess. However, the indications are that the
degree of enrichment is not as severe in West Bengal as
in the worst-affected (SE) districts of Bangladesh (e.g.
Dhar et al. 1997). Certainly, the overall areal extent of
enrichment in West Bengal is less than in Bangladesh.
The As-affected groundwaters in the Bengal Basin are
associated with sediments having total As concentrations in the range < 2–20 mg kg 1, i.e. not exceptional
by world-average values. This is not surprising given the
scale of the problem. These sediments are derived from
the drainage systems of 3 major rivers (Ganges, Brahmaputra and Meghna) which are themselves sourced
from a wide area of the Himalaya. Therefore, while it
could be argued that the source of much of the As in the
Bengal Basin sediments is derived from specific mineralised areas in the source region, these are likely to be so
widespread as to be academic and of little practical
relevance.
Isotopic evidence suggests that some groundwater
from the Bengal Basin has had a long residence time in
the aquifers. In Bangladesh, BGS and DPHE (2001)
found that 3H, an indicator of modern groundwater,
was usually detectable at a few TU in the shallowest
groundwaters but deeper groundwaters usually had
lower concentrations, typically <0.4 TU. Such low
concentrations are indicative of older groundwater, with
a large proportion having been recharged prior to the
1960s. Radiocarbon data also suggest the presence of
old groundwater at some sites. Groundwater from 10–
40 m depth in groundwaters from a study area in western Bangladesh was ‘modern’ since it contained 83%
modern C (pmc) or greater, indicating an age of the
order of decades. Shallow groundwaters collected from
south-central Bangladesh were also modern (78 pmc or
greater), although groundwater from 150 m was notably
older (51 pmc) with a model age of about 2 ka. Deep
groundwaters analysed from southern Bangladesh were
even older with 14C activities of 28 pmc or less, suggesting the presence of palaeowaters with ages of the
order of 2–12 ka.
The reasons for the distinction between groundwater
As concentrations in the shallow and deep aquifers of
the Bengal Basin are not yet well-understood. Differences between the sediments at depth may occur in
terms of absolute As concentrations and in the oxidation states and binding properties of the As to the sediments. However, it is also possible that the history of
groundwater movement and aquifer flushing in the
Bengal Basin has been important in generating the differences in dissolved As concentrations between the
shallow and deep aquifers. Older, deeper sediments have
been subject to longer periods of groundwater flow,
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
aided by greater hydraulic heads during the Pleistocene
period when glacial sea levels around the Bangladesh
landmass were up to 130 m lower than today (e.g.
Umitsu, 1993). Flushing of the deeper older aquifers
with groundwater is therefore likely to have been much
more extensive than in the Holocene sediments deposited during the last 5–10 ka. Hence, much of the As in
the deep sediments may have previously been flushed
away. Salinity becomes a problem in the near-coastal
parts of the aquifers in southern Bangladesh as a result
of saline intrusion (BGS and DPHE, 2001). This affects
the usability of the shallow aquifer in parts of Bangladesh and means that deep wells, often more than 200 m
deep, may need to be constructed to obtain fresh water.
As mentioned above, these almost always have low As
concentrations. Other potential sources of low-As water in
the region are surface water (rivers and ponds), rainwater
and dug-well water.
5.2.2. Taiwan
Arsenic problems in groundwaters in Taiwan were
first recognised during the 1960s (e.g. Tseng et al., 1968)
and chronic As-related health problems have been welldocumented by several workers since then (e.g. Chen et
al., 1985). Taiwan is the classic area for the identification of blackfoot disease but a number of other typical
health problems, including internal cancers, have been
described. High As concentrations in groundwater have
been recognised in both the SW (Tseng et al., 1968) and
NE (Hsu et al., 1997) of the island.
Kuo (1968) observed As concentrations in groundwater samples from SW Taiwan ranging between 10 and
1800 mg l 1 (mean 500 mg l 1, n=126) and found that
half the samples analysed had concentrations between
400 and 700 mg l 1. A large study carried out by the
Taiwan Provincial Institute of Environmental Sanitation established that 119 townships in the affected area
had As concentrations in groundwater of >50 mg l 1
and 58 townships had > 350 mg l 1 (Lo et al., 1977). In
north-eastern Taiwan, Hsu et al. (1997) found As concentrations in some groundwaters exceeding 600 mg l 1,
with an average of 135 mg l 1 (377 samples).
In the SW, the high As concentrations are found in
groundwaters from deep artesian wells (mostly 100–280
m) abstracted from sediments which include fine sands,
muds and black shale (Tseng et al., 1968). Groundwaters abstracted from the NE of Taiwan are also artesian, but of shallower depth (typically in the range 16–
40 m; Hsu et al., 1997). In each area, the groundwaters
are likely to be strongly reducing, a hypothesis supported by the observation that the As is present largely
as As(III) (Chen et al., 1994). However, both the geochemistry of the groundwaters and the mineral sources
in Taiwan are poorly defined at present. Groundwater
samples taken from shallow open dug wells are observed
to have low As concentrations (Guo et al., 1994).
547
5.2.3. Northern China
Arsenic has been found at high concentrations (in
excess of the Chinese national standard of 50 mg l 1) in
groundwaters from Inner Mongolia as well as Xinjiang
and Shanxi Provinces (Fig. 2; Wang, 1984; Wang and
Huang, 1994; Niu et al., 1997; Smedley et al., 2001a).
The first cases of As poisoning were recognised in Xinjiang Province in the early 1980s. Wang (1984) found As
concentrations up to 1200 mg l 1 in groundwaters from
the province. Wang and Huang (1994) reported As
concentrations of between 40 and 750 mg l 1 in deep
artesian groundwater from the Dzungaria Basin on the
north side of the Tianshan Mountains (from Aibi Lake
in the west to Mamas River in the east, a stretch of ca.
250 km). Arsenic concentrations in artesian groundwater from deep boreholes (up to 660 m) were found to
increase with depth. Shallow (non-artesian) groundwaters had observed As concentrations between <10 mg
l 1 (the detection limit) and 68 mg l 1. The concentration of As in the saline Aibi Lake was reported as 175
mg l 1, while local rivers had concentrations of between
10 and 30 mg l 1. Artesian groundwater has been used
for drinking in the region since the 1960s and chronic
health problems have been identified as a result (Wang
and Huang, 1994).
In Inner Mongolia, concentrations of As in excess of
50 mg l 1 have been identified in groundwaters from
aquifers in the Ba Men region and the Tumet Plain,
which includes the Huhhot Basin (e.g. Luo et al., 1997;
Ma et al., 1999). These areas include the towns of Boutou and Togto. In the Huhhot Basin, the problem is
found in groundwaters from Holocene alluvial and
lacustrine aquifers under highly reducing conditions and
is worst in the lowest-lying parts of the basin (Smedley
et al., 2001a). Concentrations up to 1500 mg l 1 have
been found in the groundwaters, with a significant proportion (60–90%) of the As being present as As(III).
Some shallow dug wells also have groundwater with
relatively high As concentrations (up to 556 mg l 1;
Smedley et al., 2001a). Shallow groundwaters in parts of
the region are saline as a result of evaporative concentration exacerbated by irrigation and many have high
F concentrations, although the F does not generally
correlate with As. In the affected region, As-related disease has been identified by Luo et al. (1997). Recognised
health effects include lung, skin and bladder cancer as well
as prevalent keratosis and skin-pigmentation problems.
5.2.4. Vietnam
The aquifers of the large deltas of the Mekong and
Red Rivers are now widely exploited for drinking water.
The total number of tubewells in Vietnam is unknown
but could be of the order of one million, with perhaps
150,000 in the Red River delta region. The majority of
these are private tubewells. The capital city, Hanoi, is
now largely dependent on groundwater for its public
548
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
water supply. The aquifers exploited are of both Holocene and Pleistocene age. In parts of the Red River delta
region, Holocene sediments form the shallow aquifer
which may be only 10–15 m deep. Where absent, older
Pleistocene sediments are exposed at the surface. Unlike
Bangladesh, even when the Holocene sediments are
present, there is not always a layer of fine silt-clay at the
surface. Normally the Holocene sediments are separated
from the underlying Pleistocene sediments by a clay
layer several metres thick, although ‘windows’ in this
clay layer exist where there is hydraulic continuity
between the Holocene and Pleistocene aquifers. The
total thickness of sediments is typically 100–200 m.
The groundwaters in the delta regions are usually
strongly reducing with high concentrations of Fe, Mn
and NH4. Much of the shallow aquifer in the Vietnamese part of the Mekong delta region is affected by
salinity and cannot be used for drinking water. Little
was known about the As concentrations in groundwater
in Vietnam until recently. UNICEF and EAWAG/CEC
(Hanoi National University) have been carrying out
extensive investigations to assess the scale of the problem. Results from Hanoi (Wegelin et al., 2000; Berg et
al., 2001) indicate that there is a significant As problem
in shallow tubewells in the city, particularly in the south.
Concentrations in the range 1–3050 mg l 1 (average 159
mg l 1) were reported by Berg et al. (2001). There
appears to be a large seasonal variation with significantly lower concentrations having been found at
most investigated sites during sampling in May (early
rainy season). This could be related to the local hydrology since there are significant interactions between the
aquifer and the adjacent Red River. Little is known
about the As concentrations in groundwater from the
middle and upper parts of the Mekong delta (and into
adjacent Cambodia and Laos) and other smaller alluvial
aquifers in Vietnam but investigations are presently
taking place.
5.2.5. Hungary and Romania
Concentrations of As above 50 mg l 1 have been
identified in groundwaters from alluvial sediments in the
southern part of the Great Hungarian Plain (Fig. 3) of
Hungary and neighbouring parts of Romania. Concentrations up to 150 mg l 1 (average 32 mg l 1, 85 samples) have been found by Varsányi et al. (1991). The
Plain, some 110,000 km2 in area, consists of a thick
sequence of subsiding Quaternary sediments. Groundwaters vary from Ca–Mg–HCO3-type in the recharge
areas of the basin margins to Na–HCO3-type in the lowlying discharge regions. Groundwaters in deep parts of
the basin (80–560 m depth) with high As concentrations
are reducing with high concentrations of Fe and NH4
and many have reported high concentrations of humic
acid (up to 20 mg l 1; Varsányi et al., 1991). The
groundwaters have highest As concentrations in the
lowest parts of the basin, where the sediment is finegrained. Gurzau and Gurzau (2001) reported As
concentrations up to 176 mg l 1 in the associated aquifers
of neighbouring Romania (Table 7).
5.3. Arid oxidising environments
5.3.1. Mexico
The Lagunera Region of north central Mexico has a
well-documented As problem in groundwater with significant resulting chronic health problems. The region is
arid and groundwater is an important resource for
potable supply. Groundwaters are predominantly oxidising with neutral to high pH. Del Razo et al. (1990)
quoted pH values for groundwaters in the range 6.3 to
8.9. They found As concentrations in the range 8 to 624
mg l 1 (average 100 mg l 1, n=128), with half the samples having concentrations greater than 50 mg l 1. They
also noted that most (> 90%) of the groundwater samples investigated had As present predominantly as
As(V). Del Razo et al. (1994) determined the average
concentration of As in drinking water from Santa Ana
town in the region as 404 mg l 1. The estimated population exposed to As in drinking water with > 50 mg l 1 is
around 400,000 in Lagunera Region (Del Razo et al.,
1990). Groundwaters from the region also have high
concentrations of F- (up to 3.7 mg l 1; Cebrián et al.,
1994).
High As concentrations have also been identified in
groundwaters from the state of Sonora in NW Mexico.
Wyatt et al. (1998) found concentrations in the range 2–
305 mg l 1 (76 samples) with highest concentrations in
groundwaters from the towns of Hermosillo, Etchojoa,
Magdalena and Caborca. The As concentrations were
also positively correlated with F. The highest observed
F concentration in the area was 7.4 mg l 1. The reasons
for the correlation in this case are uncertain. It is also
believed that high-As groundwaters have been found in
other parts of northern Mexico.
5.3.2. Chile
Health problems related to As in drinking water were
first recognised in northern Chile in 1962. Typical
symptoms included skin-pigmentation changes, keratosis, squamous-cell carcinoma (skin cancer), cardiovascular problems and respiratory disease (Zaldivar,
1974). More recently, As ingestion has been linked to
lung and bladder cancer. It has been estimated that
around 7% of all deaths occurring in Antofagasta
between 1989 and 1993 were due to past exposure to As
in drinking water at concentrations of the order of 500
mg l 1 (Smith et al., 1998). Since exposure was chiefly in
the period 1955–1970, this pointed to a long latency
period of cancer mortality. Other reported symptoms
include impaired resistance to viral infection and lip
herpes (Karcher et al., 1999).
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
High As concentrations have been reported in surface
waters and groundwaters from Administrative Region
II (incorporating the cities of Antofagasta, Calama and
Tocopilla) of northern Chile (Cáceres et al., 1992). The
region is arid (Atacama Desert) and water resources are
limited. High As concentrations are accompanied by
high salinity and high B concentrations. This in part
relates to evaporation but is also thought to be significantly affected by geothermal inputs from the El
Tatio geothermal field. Arsenic concentrations below
100 mg l 1 in surface waters and groundwaters are
apparently quite rare, and concentrations up to 21,000
mg l 1 have been found. Karcher et al. (1999) quoted
ranges of 100 mg l 1 to 1000 mg l 1 in raw waters (average
440 mg l 1). The affected waters of Chile are taken to be
predominantly oxidising (with dissolved O2 present),
largely because the As is reported to be present in the
waters as arsenate (Thornton and Farago, 1997 and
references cited therein). However, the geochemistry of
the aquifers of Chile is as yet poorly understood. The
aquifers are composed of volcanic rocks and sediments,
but the As sources are not well-characterised. In Antofagasta, concentrations of As in the sediments are ca. 3.2
mg kg 1 (Cáceres et al., 1992). Additional As exposure
from smelting of Cu ore has also been noted in northern
Chile (Cáceres et al., 1992).
Arsenic treatment plants were installed in the towns
of Antofagasta and Calama in 1969 to mitigate the
problems. Today, the urban populations of the major
towns are supplied with treated water from the Rivers
Toconce and Loa (Karcher et al., 1999) which is transported from the foot of the Andes mountains to the
treatment works. However, rural communities still largely
rely on untreated water supplies which contain As.
5.3.3. Argentina
The Chaco-Pampean Plain of central Argentina constitutes perhaps one of the largest regions of high-As
groundwaters known, covering around 1106 km2.
High concentrations of As have been documented from
Córdoba, La Pampa, Santa Fe, Buenos Aires and
Tucumán Provinces in particular. Symptoms typical of
chronic As poisoning, including skin lesions and some
internal cancers, have been recorded in these areas (e.g.
Hopenhayn-Rich et al., 1996). The climate is temperate
with increasing aridity towards the west. The high-As
groundwaters are from Quaternary deposits of loess
(mainly silt) with intermixed rhyolitic or dacitic volcanic
ash (Nicolli et al., 1989; Smedley et al., 1998, 2002). The
sediments display abundant evidence of post-depositional diagenetic changes under semi-arid conditions,
with common occurrences of calcrete in the form of
cements, nodules and discrete layers, sometimes many
centimetres thick.
Nicolli et al. (1989) found As concentrations in
groundwaters from Córdoba in the range 6–11500 mg
549
l 1 (median 255 mg l 1). Nicolli and Merino (2002) in a
study of the Carcarañá River Basin (Córdoba and Santa
Fe Provinces) found concentrations in the range < 10–
720 mg l 1 (mean 201 mg l 1). Smedley et al. (1998,
2002) found concentrations for groundwaters in La
Pampa Province in the range <4–5280 mg l 1 (median
145 mg l 1). Nicolli et al. (2001) found concentrations in
groundwaters from Tucumán Province of 12–1660 mg
l 1 (median 46 mg l 1). The groundwaters often have
high salinity and the As concentrations are generally
well-correlated with other anion and oxyanion elements
(F, V, HCO3, B, Mo). They are also predominantly
oxidising with low dissolved Fe and Mn concentrations.
Under the arid conditions, silicate and carbonate
weathering reactions are pronounced and the groundwaters often have high pH values. Smedley et al. (2002)
found pH values typically of 7.0–8.7; Nicolli et al.
(2001) found pH values of 6.3–9.2. Arsenic is dominantly present as As(V) (Smedley et al., 2002). Metal
oxides in the sediments (especially Fe and Mn oxides
and hydroxides) are thought to be the main source of
dissolved As, caused by desorption under high-pH conditions (Smedley et al., 1998, 2002), although the direct
dissolution of volcanic glass has also been cited as a
potential source (Nicolli et al., 1989).
5.4. Mixed oxidising and reducing environments
5.4.1. South-western USA
Many areas have been identified in the USA with As
problems in groundwater. Most of the worst-affected
and best-documented cases occur in the south-western
states (Nevada, California, Arizona). However, within
the last decade, aquifers in Maine, Michigan, Minnesota,
South Dakota, Oklahoma and Wisconsin have been
found with concentrations of As exceeding 10 mg l 1 and
smaller areas of high As groundwaters have been found
in many other States. Much water analysis and research
has been carried out in the USA, particularly in view of
the long-proposed reduction in the US-EPA drinkingwater maximum contaminant level and public concern
over the possible long-term health effects. Occurrences
in groundwater are therefore noted to be widespread,
although of those reported, relatively few have significant numbers of sources with concentrations greater
than 50 mg l 1. A recent review of the analyses of some
17,000 water analyses from the USA concluded that
around 40% exceeded 1 mg l 1 and about 5% exceeded
20 mg l 1 (percentage above 50 mg l 1 unknown; Welch
et al., 1999). The As is thought to be derived from various sources, including natural dissolution/desorption
reactions, geothermal water and mining activity. The
natural occurrences of As in groundwater occur under
both reducing and oxidising conditions in different
areas. Concentration by evaporation is thought to be an
important process in the more arid areas.
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In Nevada, at least 1000 private wells have been
found to contain As concentrations in excess of 50 mg
l 1 (Fontaine, 1994). The city of Fallon, Nevada
(population 8000) is served by a groundwater supply
with an As concentration of 100 mg l 1 which for many
years has been supplied without treatment other than
chlorination. Welch and Lico (1998) reported high As
concentrations, often exceeding 100 mg l 1 but with
extremes up to 2600 mg l 1, in shallow groundwaters
from the southern Carson Desert. These are largely
present under reducing conditions, having low dissolved-O2 concentrations and high concentrations of
dissolved organic C, Mn and Fe. The groundwaters also
have associated high pH (> 8) and high concentrations
of P (locally >4 mg l 1) and U ( > 100 mg l 1; Welch
and Lico, 1998). The high As and U concentrations
were thought to be due to evaporative concentration of
groundwater, combined with the influence of redox and
desorption processes involving metal oxides.
In groundwaters from the Tulare Basin of the San
Joaquin Valley, California, a large range of groundwater As concentrations from < 1 mg l 1 to 2600 mg l 1
have been found (Fujii and Swain, 1995). Redox conditions in the aquifers appear to be highly variable and
high As concentrations are found in both oxidising and
reducing conditions. The proportion of As present as
As(III) increases in the groundwaters with increasing
well depth. The groundwaters from the Basin are often
strongly affected by evaporative concentration with
resulting high TDS values. Many also have high concentrations of Se (up to 1000 mg l 1), U (up to 5400 mg
l 1), B ( up to 73,000 mg l 1) and Mo (up to 15,000 mg
l 1; Fujii and Swain, 1995).
Robertson (1989) also noted the occurrence of high
As concentrations in groundwaters under oxidising
conditions in alluvial aquifers in the Basin and Range
Province in Arizona. Dissolved O2 values of the
groundwaters were in the range 3–7 mg l 1. Arsenic in
the groundwater was found from a limited number of
samples to be present predominantly as As(V). The dissolved As correlated well with Mo, Se, V, F and pH, the
latter being in the range 6.9–9.3. Of the 467 samples
analysed, 7% had As concentrations greater than 50 mg
l 1. Arsenic concentrations in the sediments ranged
between 2 and 88 mg kg 1. Oxidising conditions (with
dissolved O2 present) were found to persist in the aquifers down to significant depths (600 m) despite significant groundwater age (up to 10 ka old). The high As
(and other oxyanion) concentrations are a feature of the
closed basins of the province.
5.5. Geothermal sources
As noted above, As associated with geothermal
waters has been reported in several parts of the world,
including hot springs from parts of the USA, Japan,
New Zealand, Chile, Iceland, Kamchatka, France and
Dominica (e.g. White et al., 1963; Welch et al., 1988;
Criaud and Fouillac, 1989). Parts of Salta and Jujuy
Provinces in NW Argentina also have thermal springs
with high As concentrations. In the USA, occurrences of
As linked to geothermal sources have been summarised
by Welch et al. (1988, 2000). Reported occurrences
include Honey Lake Basin, California (As up to 2600 mg
l 1), Coso Hot Springs, California (up to 7500 mg l 1),
Imperial Valley, California (up to 15,000 mg l 1), Long
Valley, California (up to 2500 mg l 1) and Steamboat
Springs, Nevada (up to 2700 mg l 1). Geothermal waters
in Yellowstone National Park also contain high concentrations of As. Thompson and Demonge (1996)
reported concentrations of < 1–7800 mg l 1 in geysers
and hot springs; values up to 2830 mg l 1 were reported
by Ball et al. (1998). These geothermal sources have given
rise to high concentrations of As (up to 370 mg l 1) in
waters of the Madison River (Nimick et al., 1998). Geothermal waters at Lassen Volcanic National Park, California have As concentrations ranging up to 27,000 mg
l 1 (Thompson et al., 1985). An As concentration of
3800 mg l 1 has also been reported from Geyser Bight,
Umnak Island, Alaska (White et al., 1963). Geothermal
inputs from Long Valley, California are believed to be
responsible for relatively high concentrations (20 mg l 1)
of As in the Los Angeles Aqueduct which provides the
water supply for the city of Los Angeles (Wilkie and Hering, 1998). Geothermal inputs also contribute significantly
to the high dissolved As concentrations (up to 20 mg l 1)
in Mono Lake, California (Maest et al., 1992).
Welch et al. (1988) noted a general relationship
between As and salinity in geothermal waters from the
USA. Despite a lack of good positive correlation
between As and Cl, geothermal waters with As greater
than ca. 1000 mg l 1 mostly had Cl concentrations of
800 mg l 1 or more. Wilkie and Hering (1998) noted the
high alkalinity and pH values (average pH 8.3) as well
as high Cl and B concentrations of As-rich geothermal
waters in Long Valley. Of 26 geothermal water samples
analysed from 5 geothermal fields in Kyushu, Japan, As
concentrations have been reported in the range 500–
4600 mg l 1. The waters are typically of Na–Cl type and
the As is present in all but one sample overwhelmingly
as As(III) (Yokoyama et al., 1993).
Robinson et al. (1995) found an As concentration in
waste geothermal brine from the main drain at the
Wairakei geothermal field in New Zealand of 3800 mg
l 1. River and lake waters receiving inputs of geothermal water from the Wairakei, Broadlands, Orakei Korako and Atiamuri geothermal fields have As
concentrations up to 121 mg l 1, although concentrations are reported to diminish significantly downstream
away from the geothermal input areas.
High As concentrations have been found in geothermal waters from the El Tatio system in the Antofagasta
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
region of Chile. The geothermal area lies in a basin (altitude 4250 m) between the volcanoes of the Andes and the
Serrania de Tucle. The geothermal waters are highly saline (Na–Cl solutions with Na concentrations in the range
2000–5000 mg l 1). Arsenic concentrations of the waters
are reported to be in the range 45,000–50,000 mg l 1 (Ellis
and Mahon, 1977). White et al. (1963) also reported As
concentrations in the range 50–120 mg l 1 for thermal
waters from Iceland and in the range 100–5900 mg l 1
for thermal waters from Kamchatka.
5.6. Sulphide mineralisation and mining-related arsenic
problems
5.6.1. Thailand
Probably the worst case of As poisoning related to
mining activity known is that of Ron Phibun District in
Nakhon Si Thammarat Province of southern Thailand.
Health problems were first recognised in the area in
1987. By the late 1990s, around 1000 people had been
diagnosed with As-related skin disorders, particularly in
and close to Ron Phibun town (Williams, 1997; Choprapawon and Rodcline, 1997). The affected area lies
within the SE Asian Tin Belt. Arsenic concentrations up
to 5000 mg l 1 have been found in shallow groundwaters
from Quaternary alluvial sediment that has been extensively dredged during Sn-mining operations. Deeper
groundwaters from an older limestone aquifer have been
found to be less contaminated (Williams et al., 1996)
although a few high As concentrations occur, presumably also as a result of contamination from the mine
workings. The mobilisation of As is believed to be caused
by oxidation of arsenopyrite, exacerbated by the former
Sn-mining activities. The recent appearance in groundwater has occurred during post-mining groundwater
rebound (Williams, 1997).
5.6.2. Ghana
Ghana is an important Au-mining country and
mining has been active since the late 19th century.
Today, Ghana produces about one third of the world’s
Au. The most important mining area is the Ashanti
Region of central Ghana. As with Ron Phibun District in
Thailand, the Au is associated with sulphide mineralisation, particularly arsenopyrite. Arsenic mobilises in the
local environment as a result of arsenopyrite oxidation,
induced (or exacerbated) by the mining activity. Around
the town of Obuasi, high As concentrations have been
noted in soils close to the mines and treatment works
(Amasa, 1975; Bowell, 1992; 1993). Some high concentrations have also been reported in river waters close
to the mining activity (Smedley et al., 1996).
Despite the presence of high As concentrations in the
contaminated soils and in bedrocks close to the mines,
Smedley et al. (1996) found that many of the groundwaters of the Obuasi area had low As concentrations,
551
with a median concentration in tubewell waters of just 2
mg l 1. Some higher concentrations were observed (up to
64 mg l 1) but these were not generally in the vicinity of
the mines or related directly to mining activity. Rather, the
higher concentrations were found to be present in relatively reducing groundwaters (Eh 220–250 mV). Oxidising
groundwaters, especially from shallow hand-dug wells,
had low As concentrations. This was taken to be due
retardation of As by adsorption onto hydrous ferric oxides under the ambient low pH condition of the groundwaters (median pH 5.4 in dug wells; 5.8 in tubewells;
Smedley, 1996; Smedley et al., 1996).
5.6.3. United States
Arsenic contamination from mining activities has
been identified in numerous areas of the USA, many of
which have been summarised by Welch et al. (1988;
1999). Groundwater from some areas has been reported
to have very high As concentrations locally (up to 48,000
mg l 1). Well-documented cases of As contamination
include the Fairbanks Au-mining district of Alaska
(Wilson and Hawkins, 1978; Welch et al., 1988), the
Coeur d’Alene Pb–Zn–Ag mining area of Idaho, (Mok
and Wai, 1990), Leviathan Mine, California (Webster et
al., 1994), Kelly Creek Valley, Nevada (Grimes et al.,
1995), Clark Fork river, Montana (Welch et al., 2000)
and Lake Oahe in South Dakota (Ficklin and Callender,
1989). Some mining areas of the USA have significant
problems with acid mine drainage resulting from extensive oxidation of Fe sulphide minerals. In these, pH
values can be extremely low and Fe oxides, produced
from the oxidation reaction, dissolve and release bound
As. Iron Mountain has some extremely acidic minedrainage waters with negative pH values and As concentrations in the mg l 1 range (Nordstrom et al., 2000).
In Wisconsin, As and other trace-element problems in
groundwater have arisen as a result of the oxidation of
sulphide minerals (pyrite and marcasite) present as a
discrete secondary cement horizon in the regional
Ordovician sandstone aquifer. Concentrations of As up
to 12,000 mg l 1 have been reported in the well waters
(Schreiber et al., 2000). The oxidation appears to have
been promoted by groundwater abstraction which has led
to the lowering of the piezometric surface at a rate of
around 0.6 m a 1 since the 1950s with partial dewatering
of the aquifer. The high As concentrations are observed
where the piezometric surface intersects, or lies close to,
the sulphide cement horizon (Schreiber et al., 2000).
5.6.4. Other areas
Many other areas have increased concentrations of
As in water, soils, sediments and vegetation as a result
of local mineralisation, exacerbated by mining activity.
Documented cases include the Lavrion region of
Greece, associated with Pb and Ag mining (Komnitsas
et al., 1995), the Zimapán Valley of Mexico (Armienta
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et al., 1997), parts of SW England (Thornton and Farago, 1997), South Africa and Zimbabwe (Jonnalagadda
and Nenzou, 1996). Although severe contamination of
the environment has often been documented in these
areas, the impact on groundwaters used for potable
supply is usually minor.
Increased concentrations of dissolved As have also
been found in parts of the world with local mineralisation which has not been mined. Boyle et al. (1998) recorded concentrations up to 580 mg l 1 in groundwaters
from an area of sulphide mineralisation in Bowen Island,
British Colombia. Heinrichs and Udluft (1999) found
often high concentrations in groundwater from the
Upper Triassic Keuper Sandstone in northern Bavaria.
Out of 500 wells, 160 had As concentrations in the range
10–150 mg l 1. The nature of the mineralisation in this
aquifer was not clearly identified. As yet unidentified
areas of mineralisation could be quite widespread,
although these are likely to be on a local scale.
6. Common features of groundwater arsenic problem
areas
6.1. A hydrogeochemical perspective
Historically, as new sources of high As groundwaters
have been found, treatment plants have been built and
the problem has receded from public attention. Although
much recent research has been carried out in the USA,
there have been few detailed hydrogeochemical and
hydrogeological studies in the As-affected aquifers from
many other parts of the world. Despite this, sufficient is
already known that it is useful to attempt to bring
together some of the common features and to speculate
about the critical factors that can lead to high-As
groundwaters. This will help to focus future scientific
studies and should provide some guidance to water
providers who have to undertake a rapid assessment of
water supplies for As. It is helpful to consider the formation of high-As groundwaters in terms of the 3 major factors involved, namely, the source of the As, its
mobilisation, and its subsequent transport (or lack of it).
6.2. The source of arsenic
In the cases where affected groundwaters are found
close to obvious geological or industrial sources rich in
As (geothermal springs, drainage from mineralised and
mining areas, specific contaminant sources), it is clear
that the anomalously high As concentrations in the
source region are responsible. The extent of this contamination is usually highly localised because the geochemical conditions within most aquifers do not favour
As mobilisation on a regional scale. Areas affected by
geothermal activity are potentially more widespread
since in this case mobilisation of As is not an issue: As is
already present in solution and the size of geothermal
reservoirs can be large. Perhaps more puzzling is the
way in which exceptionally high concentrations of As,
up to several mg l 1, are found in groundwaters from
areas with apparently near-average source rocks. In
aquifers with extensive high-As groundwater, this
appears to be the rule rather than the exception. Most of
these cases arise in aquifers derived from relatively
young aquifer materials, often consisting of alluvium or
loess where the total As concentrations in the sediments
are usually in the range 1–20 mg kg 1. Recognition of
this fact is a recent development and its late appreciation has delayed the discovery of many high-As
groundwater provinces.
A critical point is that the drinking-water limit for As
is very low in relation to the overall abundance of As in
the natural environment. Fortunately, most of this As is
normally immobilised by various minerals, particularly
Fe oxides, and so is not available for abstraction. However, it only takes a small percentage of this ‘solid’ As to
dissolve or desorb to give rise to a serious groundwater
problem. This can provide an explanation for both the
oxidising and reducing high-As environments. An
abundant source of Fe oxides with its surface-bound
and coprecipitated As provides a ready source of As
that may be released given an appropriate change in
geochemical conditions.
6.3. Arsenic mobilisation—the necessary geochemical
trigger
There appear to be two key factors involved in the
formation of high-As groundwaters on a regional scale:
firstly, there must be some form of geochemical trigger
which releases As from the aquifer solid phase into the
groundwater. Secondly, the released As must remain in
the groundwater and not be flushed away. There are a
number of possible geochemical triggers. In mining and
mineralised areas, oxidation of sulphide ores may be
triggered by influxes of O2 or other oxidising agents.
However, in most As-affected aquifers, the most important trigger appears to be the desorption/dissolution of
As from oxide minerals, particularly Fe oxides. An
important feature of this process is that the initial
adjustment is probably quite rapid since it involves a
shift from one point on the adsorption isotherm to
another and adsorption reactions, being surface reactions, are usually rapid. The rate-limiting factors are
probably those controlling the major changes in pH, Eh
and associated water quality parameters of the aquifer.
These are in part related to physical factors such as the
rate of diffusion of gases through the sediment and the
rate of sedimentation, in part due to the extent of
microbiological activity and in part related to the rates
of chemical reactions. However, many of these factors
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
553
are likely to be rapid on a geological time scale (tens of
thousands of years and longer). Dissolution reactions
are slower but even oxide dissolution is rapid on a geological time scale and can be observed in a matter of
weeks or even days in flooded soils (Ponnamperuma et
al., 1967; Masscheleyn et al., 1991).
A qualification is that if diagenetic changes to the
oxide mineral structure are important (see below) or if
burial of sediment is important, then there could be a
slow release of As over a much longer time scale. Details
of the rate of release of As and how this varies with time
are not yet clear. It is likely that the rate will diminish
with time with the greatest changes occurring in the
early stages. Natural groundwater flushing means that
very slow releases of As are likely to be of little consequence since the As released will not tend to accumulate to a significant extent. A corollary of this hypothesis
is that once the diagenetic readjustment has taken place
and the sediments have equilibrated with their new
environment, there should be little further release of As.
This contrasts with some mineral-weathering reactions
which occur in ‘open’ systems and can continue for
millions of years until all of the mineral has dissolved.
Seen in this context, the desorption/dissolution of As
from metal oxides in young aquifers is essentially a step
change responding to a new set of conditions. As discussed above, the type of reactions that may occur can
be seen today most clearly where they occur at a small
spatial scale and over a short time scale, for example,
across a redox boundary in a lake sediment. The geochemical triggers involved could arise for a number of
possible reasons. Below the authors speculate what
these might be. Some model calculations of their possible
impact are given in BGS and DPHE (2001).
at pH 7 with a river water containing 1 mg l 1 of As and
0.05 mol l 1 NaCl, then according to the DLM, HFO
has an As loading of almost 15,000 mg kg 1. This is
equivalent to 23 mg As kg 1 sediment. If the pH is then
raised while maintaining a constant total As concentration in the sediment plus porewater, equivalent to evolution within a ‘closed’ system, then desorption of As
occurs (Fig. 5). There is a sharp increase in porewater
As concentration above pH 8.5. Above pH 9, the As(V)
concentration can exceed 1000 mg l 1. An ionic strength
of 0.05 M is greater than found in most river waters but
is useful for maintaining a constant ionic strength in the
calculations. Normally the effects of changes in ionic
strength are relatively small. This system assumes no
competing anions in solution.
More realistically, other anions (e.g. phosphate, bicarbonate and silicate) in the river water will compete for
sorption sites on the HFO and thereby reduce the initial
As loading. If we assume for example that the river water
also contains 10 mg l 1 phosphate-P, then the initial As
loading is reduced to 5600 mg kg 1 HFO. However, while
this reduces the As release at high pH to some extent, the
pH effect is still strong (Fig. 5). Phosphate (and other
adsorbed anions) are also released along with As at high
pH. At pH 9, the calculated total P concentration in the
above example is 13 mg l 1.
There are several reasons why the pH might increase
but the most important in the present context is the
uptake of protons by mineral weathering and ionexchange reactions, combined with the effect of evaporation in arid and semi-arid regions. This pH increase
is commonly associated with the development of salinity
and the salinisation of soils. Inputs of high-pH geothermal waters may be important in maintaining high As
6.3.1. Desorption at high pH under oxidising conditions
Under the aerobic and acidic to near-neutral conditions typical of many natural environments, As is very
strongly adsorbed by oxide minerals as the arsenate ion.
The highly non-linear nature of the adsorption isotherm
for arsenate ensures that the amount of As adsorbed is
relatively large, even when dissolved concentrations of
As are low. Adsorption protects many natural environments from widespread As toxicity problems. As pH
increases, especially above pH 8.5, As desorbs from the
oxide surfaces, thereby increasing the concentration of As
in solution. The impact of this is magnified by the high
solid/solution ratios typical of aquifers (3–10 kg l 1).
It is possible to demonstrate this effect with some
model calculations. The Dzombak and Morel (1990)
diffuse double layer model (DLM) for hydrous ferric
oxide (HFO), and its accompanying thermodynamic
database, provides a simple means of calculating the
amount of As adsorbed under various conditions. If we
assume that a sandy sediment with 25% porosity contains
1 g Fe kg 1 as HFO and that it is initially in equilibrium
Fig. 5. Calculated increase in As concentration when the pH of
a sediment containing 1 g kg 1 Fe as HFO is increased from its
initial value of pH 7 under closed-system conditions.
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P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
concentrations in some alkaline lakes. Desorption at
high pH is the most likely mechanism for the development of groundwater-As problems under the oxidising
conditions and would account for the observed positive
correlation of As concentrations with increasing pH
(e.g. Robertson, 1989; Smedley et al., 2002).
As such a pH increase induces the desorption of a
wide variety of oxyanions, other oxyanions such as
phosphate, vanadate, uranyl and molybdate will also
tend to accumulate. There is evidence that this is indeed
the case (Smedley et al., 2002). These specificallyadsorbed anions all interact with the adsorption sites on
the oxides in a competitive way and so influence, in a
complex way, the extent of binding of each other. This is
not well understood in a quantitative sense. Phosphate
in particular may play an important role in As binding
since it is invariably more abundant than As, often by a
factor of 50 or more (in molar terms), and is also
strongly bound to oxide surfaces. At pH 7, arsenate is as
strongly sorbed as phosphate (Hingston et al., 1971).
The role of HCO3 , often the major anion in As-affected
groundwaters, in promoting the desorption of arsenate
is unclear at present. Wilkie and Hering (1996) did not
find any large effects of 1 mM HCO3 on As(V) or
As(III) sorption by HFO at pH 6 and 9 (but greatest for
As(III) at pH 9). Recent studies have shown that a high
pCO2 (and implicitly HCO3 concentration) significantly
reduces the binding of chromate, another strongly sorbed anion, by goethite (Villalobos et al., 2001). Bicarbonate concentrations in high-As groundwaters usually
exceed 300 mg l 1 (5 mM) (Smedley et al., 1998; 2001a
BGS and DPHE, 2001) and can exceed 1000 mg l 1 (16
mM). Hence, experiments need to be extended to these
high concentrations to verify the role of HCO3 in sorption
processes.
The role of dissolved organic C (fulvic and humic
acids) is also uncertain, at least from a quantitative
point of view. Humic substances have been shown to
reduce As(III) and As(V) sorption by Fe oxides under
some conditions (Xu et al., 1991; Bowell, 1994) and
high-As groundwaters are associated with high humicacid concentrations in some aquifers (Smedley et al.,
2001a). However, direct evidence for a causal link
between dissolved organic C and As desorption does not
yet exist.
Some cations, because of their positive charge, may
promote the adsorption of negatively charged arsenate
(Wilkie and Hering, 1996). Calcium and Mg are likely to
be the most important cations in this respect because of
their abundance in most natural waters and their +2
charge. Divalent Fe may be important in reduced waters
and Al in acidic waters. Silica also exerts a control on
the sorption of As (Swedlund and Webster, 1998).
The aridity described above enables the high pH
values to be maintained and minimises the flushing of
the As. It also allows the build-up of high Cl and F
concentrations. Other high-pH environments (up to pH
8.3), particularly open-system calcareous environments,
are likely to be too well flushed to allow any released As
to have accumulated. Since calcium arsenate is highly
insoluble, it is also likely that arsenate will be strongly
sorbed by calcium carbonate minerals, although this has
not been demonstrated.
The pH dependence of adsorption is critical but has
not yet been measured in detail for any aquifer materials, especially in the presence of typical groundwater
compositions. The pH dependence is likely to depend to
some extent on the heterogeneity of the aquifer material.
Other specifically adsorbed anions, particularly phosphate and perhaps HCO3, may also significantly affect
the pH dependence of As(V) and As(III) binding. High
pH values cannot explain the development of high As
concentrations in reducing environments such as Bangladesh since groundwaters in reducing environments
normally have a near-neutral pH.
6.3.2. Arsenic desorption and dissolution due to a change
to reducing conditions
The onset of strongly reducing conditions, sufficient
to enable Fe(III) and probably SO4 reduction to take
place, appears to be another trigger for the release of
As. The most common cause of this is the rapid accumulation and burial of sediments. This occurs in river
valleys, especially in broad valleys with wide meandering river channels carrying heavy sediment loads.
Large, rapidly advancing deltas are an extreme case.
The organic C content of the buried sediment will largely determine the rate at which reducing conditions are
created. Freshly-produced soil organic matter readily
decomposes and it does not take much of this to use up
all of the dissolved O2, NO3 and SO4. Reducing conditions can only be maintained if the diffusion and convection of dissolved O2 and other oxidants from the surface is
less rapid than their consumption. This is helped if there is
a confining layer of fine-grained material close to the
surface. This often occurs in large deltas where finegrained overbank deposits overlie coarser-grained alluvial
deposits.
While the detailed reactions (in surface chemical
terms) that occur when reduction takes place are not
well understood, the change from normally strongly
adsorbed As(V) to normally less strongly adsorbed
As(III) may be one of the first reactions to take place,
although not all the evidence supports this (e.g. De Vitre
et al., 1991). A change in the redox state of the adsorbed
ions could have wider-ranging repercussions since it will
also affect the extent to which other anions can compete
for adsorption sites. Phosphate-arsenite competition, for
example, is likely to be less important than phosphatearsenate competition. There is also the potential for
arsenite-arsenate competition. Model calculations suggest that adsorbed phosphate can reverse the relative
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
555
affinity of As(III) and As(V) at near-neutral pH values
(BGS and DPHE, 2001). These complexities are poorly
understood at present but are important if reliable
quantitative predictions of As concentrations under
reducing conditions are to be made.
In the absence of other specifically adsorbed anions,
As(III) adsorption by Fe oxides is practically independent of pH. This contrasts sharply with that of As(V) as
described above. In the ‘real world’, some pH dependence in As(III) adsorption will be induced by secondary interactions arising from the pH dependence of
other specifically adsorbed and competing ions such as
phosphate and even As(V).
6.3.3. Reduction in surface area of oxide minerals
Disordered and fine-grained Fe oxides, which may
include HFO, lepidocrocite, schwertmannite and magnetite, are commonly formed in the early stages of
weathering. Freshly-precipitated HFO is extremely finegrained with cluster sizes of about 5 nm diameter and a
specific surface area of 600 m2 g 1 or greater. HFO
gradually transforms (crystallises) to more ordered
forms such as goethite or hematite with correspondingly
large crystal sizes and reduced surface areas. This ‘ageing’ reaction can take place rapidly in the laboratory but
the rate in nature is likely to be somewhat inhibited by
the presence of other ions, particularly by strongly
adsorbed ions such as Al, PO4, SO4, AsO4, HCO3 and
silicate (Cornell and Schwertmann, 1996).
One consequence of this reduction in surface area is
that the amount of As(V) adsorption may decrease on a
weight for weight basis. If the site density (site nm 2)
and binding affinities of the adsorbed ions remain constant, then as the specific surface area of an oxide
mineral is reduced, some of the adsorbed ions will be
desorbed. The specific surface area of HFO on which
the DLM was calibrated by Dzombak and Morel (1990)
was taken as 600 m2 g 1. Aged products, such as goethite typically have specific surface areas of 150 m2 g 1
or less and even less for hematite (Cornell and Schwertmann, 1996).
Assuming that the site density remains unchanged
during ageing and the binding affinity of As(V) and
protons also remains unchanged, it is possible to calculate the effect that a reduction in surface area might
have. If we assume similar initial conditions to the pH
example given above, i.e. HFO in equilibrium with a
solution containing 1 mg As(V) l 1 and a sediment with
25% porosity and a HFO concentration equivalent to
1g Fe kg 1, then these calculations show that there is a
marked increase in As concentration in solution when
the surface area is reduced to below 300 m2 g 1 (Fig. 6).
Assuming that the HFO was also initially in equilibrium with 10 mg l 1 phosphate-P increases the effect
somewhat. These calculations are strongly dependent on
the shape of the adsorption isotherm and the initial As
Fig. 6. Calculated increase in As concentration when the specific surface area of HFO in a sediment containing 1 g kg 1 Fe
as HFO is reduced from its initial value of 600 m2 g 1 under
closed-system conditions.
loading, and are therefore sensitive to all the interactions that control the isotherm. For example, for As(III)
which has a close-to-linear isotherm at typical groundwater concentrations, the effect of a reduction in surface
area is less dramatic and is closer to a linear function of
the surface-area reduction.
6.3.4. Reduction in binding strength between arsenic and
mineral surfaces
While the calculations outlined above are useful in the
sense that they give some indication of possible surface
area effects, they make important and unverified assumptions, and are almost certainly an oversimplification of
reality. For example, some of the desorbed ions are likely
to be incorporated into the evolving oxide structure as a
solid solution. This will reduce the amount of As
released. Also, if the surface structure changes, then it is
likely that the binding affinity of arsenate ions and protons will also change since the binding affinity is closely
related to surface structure. For example, Kinniburgh et
al. (1977) found that while aged HFO gels had a significantly reduced proton buffer capacity, and by implication a reduced specific surface area, the binding of
trace Zn2+ increased on ageing, implying that the binding affinity of Zn increased with increasing crystallisation
of the gel and that this more than compensated for the
reduction in specific surface area, i.e. the isotherm changed from a low-affinity isotherm to a high-affinity isotherm.
Under strongly reducing conditions, it appears that
additional processes could operate which may lead to a
reduction in the overall adsorption of As. Specifically
for Fe oxides, some of the surface Fe could be reduced
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from Fe(III) to Fe(II) to produce a mixed-valence oxide
perhaps akin to that of a magnetite or a green rust. This
would tend to reduce the net positive charge of the surface (or increase its net negative charge) and would
thereby reduce the electrostatic interaction between the
surface and anions. This could result in the desorption
of As and a corresponding large increase in the concentration of As in solution (BGS and DPHE, 2001).
On balance, laboratory and field evidence suggests
that at micromolar As concentrations, freshly-formed
HFO binds more As than goethite on a mole of Fe basis
(De Vitre et al., 1991) and so a reduction in affinity
appears to be more probable. In Bangladesh, areas with
high-As groundwaters tended to correspond with those
areas in which the sediments contain a relatively high
concentration of oxalate-extractable Fe (BGS and
DPHE, 2001). This provides indirect support for the
importance of Fe oxides. Some early work (Khalid et
al., 1977) also showed that the sorption of phosphate by
reduced soils was less than that for their oxidised
equivalents but that this was only true at low (near natural) phosphate loadings. It can be speculated that the
soils and sediments most sensitive to arsenic release on
reduction and ageing are those in which Fe oxides are
abundant, in which HFO is a major fraction of the Fe
oxides present, and in which other As-sorbing minerals
are relatively scarce. While the existing evidence is
somewhat contradictory, it tends to suggest that a
change from aerobic to anaerobic conditions often
results in a net release of As.
6.3.5. Mineral dissolution
Mineral dissolution reactions tend to be most rapid
under extremes of pH and Eh. Iron oxides dissolve
under strongly acidic conditions and under strongly
reducing conditions. Minor elements, including As,
present either as adsorbed (labile) As or as irreversiblybound (non-labile) As will also tend to be released during this dissolution. This can explain, in part at least, the
presence of high As concentrations in acid mine drainage and in strongly reducing groundwaters. The
reductive dissolution of Fe(III) oxides has been extensively discussed in the context of estuarine, lake and
river particles and sediments (Davison, 1993; Lovley
and Chapelle, 1995). It accounts for the high Fe(II)
content of anaerobic waters, and the ubiquity of Fe in
the environment means that waters with near-neutral
pH and high Fe are invariably anaerobic waters. While
this process undoubtedly accounts for some of the As
found in reducing groundwaters, it does not appear to
be sufficient to account for all, or even most, of the As
released. Congruent dissolution of a typical Fe oxide
without any precipitation of secondary Fe phases would
only release a few, or a few tens, of mg As l 1 which is
far less that that found in many As-rich reducing
groundwaters. The precise sequence of events that
occurs during the reductive dissolution of Fe(III) oxides
containing adsorbed and coprecipitated As is complex
and has not been studied in detail. It is even more complex than for phosphate release because of the variable
oxidation state of As. The reduction of both the Fe(III)
oxide and the As(V) are microbially catalysed and the
relative rates depend on the viability and nutrient supply
of the specific microbial strains involved (Ehrlich, 1996).
Reductive dissolution cannot of course explain the
occurrence of high-As oxidising groundwaters.
Manganese oxides also undergo reductive desorption
and dissolution and so could contribute to the As load
of groundwaters in the same way as Fe. Certainly many
of the reducing groundwaters of Bangladesh contain
high concentrations of Mn (DPHE/BGS/MML, 1999;
BGS and DPHE, 2001). The Mn oxide surfaces also
readily catalyse the oxidation of As(III) (Oscarson et al.,
1981). It is not known whether or not the dissolution of
carbonate minerals (calcite, dolomite, siderite), which
are common minerals in aquifers, contribute significantly to the release of As to groundwater, or its
uptake from groundwater.
Sulphide oxidation, particularly pyrite oxidation, can
also be an important source of As especially where these
minerals are freshly exposed as a result of a lowering of
the water table. This can occur locally in and around
mines and on a more regional scale in aquifers. In extreme
cases, this can lead to highly acidic groundwaters rich in
SO4, Fe and trace metals. As the dissolved Fe is neutralised, it tends to precipitate as a hydrous ferric oxide
(sometimes schwertmannite) with a resultant adsorption
and coprecipitation of dissolved As(V). In this sense,
pyrite oxidation is not a very efficient mechanism for
releasing As to surface and groundwaters.
6.4. Transport—historical groundwater flows
The geochemical triggers described above may release
As into groundwater but are not alone sufficient to
account for the distribution of high-As groundwaters
observed. The additional factor is that the released As
must not have been flushed away or diluted by normal
groundwater flow. This also places a time dimension on
the problem since the rate of release must be set against
the accumulated flushing of the aquifer that has taken
place during the period of release. The rocks of most
aquifers used for drinking water are several hundred
million years old and yet contain groundwater that may
be only a few thousand years old or younger. This
implies that many pore volumes of fresh water have
passed through the aquifer over its history. The oldest
fresh groundwaters are found in the Great Artesian
Basin of Australia, being up to about 400 ka (Collon et
al., 2000). The water moves slowly through this aquifer
and over its 2.5 Ga history, there have been many pore
volumes of fresh water flushed through the system. Any
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
desorbed As will have long since disappeared. The same
is true of most young aquifers with actively flowing
groundwater. On the other hand, many deltaic and
alluvial aquifers are characterised by relatively young
sediments and often relatively old groundwater. The
relative ages of the aquifer rocks and of the groundwater
are important. It is only when the geochemical trigger to
mobilise As and the hydrogeological regime to preserve it
are both operating that we see high groundwater As
concentrations on a regional scale.
It is also necessary to consider historical groundwater
flows which may have been very different from the present-day flows. One of the more significant ‘recent’
events is the global change in sea levels that has occurred over the last 130 ka (Pirazzoli, 1996). Sea levels
reflect global climate patterns. Between about 120 ka
ago and 18 ka ago, the sea level steadily declined (with a
few fluctuations) as glaciers expanded. The last glaciation was at a maximum some 21–13.5 ka ago with sea
levels being up to 130 m below present mean sea level.
This was a worldwide phenomenon and would have
affected all then existing coastal aquifers. Continental
and closed basin aquifers on the other hand would have
been unaffected. The hydraulic gradient in coastal aquifers would therefore have been much greater than at
present which would have resulted in correspondingly
large groundwater flows and extensive flushing. The As
in these older aquifers would therefore tend to have
been flushed away. The deep unsaturated zone would
also have led to more extensive oxidation of the shallower horizons with possible increased sorption of As to
Fe(III) oxides. Relics of these high flows are seen in the
extensive fissure formation in some of the world’s older
carbonate aquifers. Between some 13.5–7 ka ago,
warming occurred and sea levels rapidly rose to their
existing levels. Therefore aquifers that are younger than
some 7 ka old will not have been subjected to this
increased flushing that occurred during the most recent
glaciation.
The time taken to flush an aquifer depends on many
factors (Appelo and Postma, 1994). A critical factor is
the number of pore volumes of ‘fresh’ water that have
passed through the aquifer since the initial release of As
has taken place. The other important factor is the partitioning of As between the aquifer solid phase and the
groundwater. This determines the ease with which As is
flushed out and is related to the slope of the adsorption
isotherm (Appelo and Postma, 1994). In simple cases,
this can be expressed by the partition coefficient, Kd.
The greater the Kd, the greater the capacity of the sediment to withstand changes and the slower the As will
tend to be flushed from the aquifer. The Kd depends on
many factors both relating to the aquifer material itself
and to the chemistry of the groundwater, i.e. its pH, As
concentration and speciation, phosphate concentration
and so on. In practice, the adsorption isotherms are
557
usually non-linear, which means that the Kd varies with
concentration. This leads to more complex transport
behaviour but the same general principles apply. In
aquifers with high-As groundwater, the Kd will be less
than under ‘normal’ oxidised conditions since a reduced
Kd is precisely the reason for the development of As
problems in aquifers. There have so far been no reliable
studies of Kd values applicable to As-affected aquifers.
The greater the quantity of As involved, the more strongly
it is adsorbed and the slower the rate of groundwater
movement, the longer that high-As groundwaters will
persist.
As described above, the number of pore volumes that
have passed through the aquifer is a function of the
groundwater flow velocity integrated over the time since
sediment burial. In Bangladesh, the age of sediment vs
depth relationship is particularly important since this
has a direct bearing on the extent of flushing. Many of
the shallow sediments in southern Bangladesh are less
than 13 ka old, even less than 5 ka old, and so will not
have experienced the extensive flushing of the last glacial
period. These sediments are from where the majority of
the tubewells abstract water. Certainly at present, flushing is slow because of the extremely small hydraulic
gradients especially in southern Bangladesh. However,
deeper and older sediments, which may exceed 13 ka
old, will have been subjected to more extensive flushing.
This may account for the ‘As-free’ groundwaters found
in the deep aquifers of Bangladesh. Geochemical factors
may also play a role since the evidence is that while the
deep groundwaters are currently reducing, they are less
strongly reducing than the shallow aquifers. Certainly,
the aquifers in the Pleistocene uplifted alluvial sediments
of the Barind and Madhupur Tracts (Fig. 4) will have
been well flushed since they are at least 25–125 ka old.
These sediments invariably yield low-As groundwaters,
typically containing less than 0.5 mg l 1 As. A complication is that the Bengal Basin is locally rapidly subsiding
and filling in with sediments. This adds to the high
degree of local and regional variation.
Regional flow patterns are not the only important
factors. At a local scale, small variations in relief or in
drainage patterns may dictate local flow patterns and
hence the distribution of As-rich groundwater. For
example, there is evidence from Argentina that the
highest groundwater As concentrations are found in the
slightly lower areas where seasonal discharge occurs
(Smedley et al., 2002). The same is true in Inner Mongolia (Smedley et al., 2001a) and may also be true in
Bangladesh. In any case, it is a characteristic of
groundwater As problem areas that there is a high
degree of local-scale variation. This reflects the poor
mixing and low rate of flushing, characteristic of the
affected aquifers.
It is clear that flat low-lying areas, particularly large
plains and delta regions, are particularly prone to
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P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
potentially high-As groundwaters since they combine
many of the risk factors identified above. The process of
delta development also favours the separation of
minerals based on particle size and produces the characteristic upwardly fining sequences of sand–silt–clay.
These lead to confining or semi-confining layers which
aid the development of strongly reducing conditions.
The youngest, distal part of the deltas will tend to contain the greatest concentration of fine-grained material
and this provides an abundant source of As in the form
of colloidal-sized oxide materials. Flocculation of colloidal material, including Fe oxides, at the freshwatersea water interface will tend to lead to relatively large
concentrations of these colloids in the lower parts of a
delta. The larger the delta, and the more rapid the
infilling, the lower the hydraulic gradient and the less
flushing that is likely to have occurred. However, some
deltas, even large deltas, may be so old and well-flushed
that even the existing low hydraulic gradients will have
been sufficient to flush away any desorbed or dissolved
As.
6.5. Future developments in arsenic research
It is necessary to know how As moves in an aquifer to
predict how concentrations might change in the future.
Arsenic, like any other solute, moves in response to the
flow of groundwater and its interaction with the aquifer
solid phase. Adsorption or precipitation reactions will
tend to retard movement relative to that of the groundwater whereas the co-transport of chemicals, including
phosphate from fertilisers, that enhance the release of
As could lead to its more rapid movement through the
aquifer, albeit limited by the rate of flow of the groundwater. Establishing the basic groundwater flow patterns
within an aquifer is a prerequisite to understanding the
movement of As. The concentration profile of a nonreactive solute such as Cl can help to establish this. Agerelated tracers such as 3H, 14C and CFCs can also help,
as well as basic hydrogeological investigations of the
aquifer.
Aside from the basic hydrogeology of the aquifer, it is
also important to understand quantitatively the solidsolution interactions that take place. This refers principally to the nature of the adsorption–desorption isotherms and the mechanisms of reductive dissolution of
Fe and Mn oxides. It is likely that what is conveniently
called ‘reductive dissolution’ is in fact a mixture of desorption, dissolution and structural rearrangement of the
oxides themselves. A two-stranded approach is required:
firstly, a detailed characterisation of sediments and
associated porewaters is needed from a variety of aquifers, both affected and not affected, akin to that undertaken by limnologists and oceanographers when
studying their sedimentary environments. In reduced
aquifers, special care should be taken to avoid oxidation
of the sediment. Secondly, these field studies need to be
backed up by new theoretical advances in modelling the
relevant surface chemical reactions of the oxides and
sediments, particularly in reducing environments. This
will involve both modelling and laboratory work.
Calculations of the rate of movement of As through
an aquifer depend on knowing the appropriate solidsolution partition coefficient (Kd), or more particularly,
on knowing the nature of the adsorption isotherm and
in being able to predict how the partitioning changes
with changes in groundwater chemistry. Therefore,
there is a need for laboratory studies of the interaction
of arsenate, and if appropriate of arsenite also, with the
affected aquifer materials. These will need to be carried
out under conditions as close as possible to those found
in the field including reducing conditions if appropriate.
This can be difficult. These studies need to be backed up
by laboratory investigations of the interaction of As
with model oxide materials to establish better models
for competitive adsorption of both arsenate and arsenite
with other common anions and cations. It is likely that
this will lead to the development of new models, or at
least to a refinement of existing ones. Any new adsorption models need to be incorporated into a groundwater
solute transport package.
Reductive dissolution of oxides with adsorbed As is
poorly understood and needs careful experimental
investigations to establish the sequence of events in terms
of changes in As and Fe speciation, changes in mineral
surface chemistry and the kinetics and stoichiometry of
Fe and As release.
6.6. Identification of ‘at risk’ aquifers
This review has outlined some of the factors that are
likely to be responsible for the development and persistence of As-rich groundwaters. These can be summarised in the form of various ‘risk’ factors that can
help to identify groundwater provinces which deserve
the highest priority for As testing (Fig. 7). These factors
relate to the nature of the environment, particularly in
relation to the source of As and the hydrogeology of the
aquifer, as well as to the possible processes that can lead
to the mobilisation of As. The best option is obviously
to test the groundwater directly for As enrichment but
where this is not possible, other water-quality data may
provide some indication of the likelihood of arsenic
enrichment. No single factor is sufficient to identify a
problem area but if collectively many of the environmental characteristics and water-quality indicators
point towards this, then some kind of testing should be
undertaken as a matter of priority. A stratified random
survey is the best approach with the stratification going
across geological or landform types. Fig. 7 can only be
used to identify susceptible provinces, not individual
wells.
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
559
Fig. 7. Classification of groundwater environments prone to As problems from natural sources. Not all the indicators of low rates of
flushing necessarily apply to all environments.
7. Concluding remarks
Much research has been carried out on As, particularly in the last few years, as a response to the growing
evidence of the element’s detrimental health effects, to
reductions in regulatory limits for As, and to the recognition of the scale of arsenic enrichment in Bangladesh
and elsewhere. This review has attempted to characterise
the distribution of As in the environment, and to
describe the main geochemical controls on its speciation
and mobilisation. The documented As-rich areas are
also summarised and the main types of environment
under which As concentrations in waters are expected to
be problematic are highlighted. Doubtless there are
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P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
other areas of the world, principally aquifers, with as yet
unrecognised problems. As widespread As testing,
health awareness and diagnosis improve internationally,
these are likely to emerge gradually.
For many water providers, including non-governmental organisations working in rural communities in
developing countries, As represents a new and poorly
understood threat. It is almost certain that there will
have been little or no As testing and there is likely to be
a general lack of understanding of the issues involved.
In some cases, such as Bangladesh, the size of the As
testing programme required is unprecedented in scale.
In other cases, there is a lack of appreciation of a
potential problem, or the lack of suitable facilities for
testing As. A rapid, randomised testing programme will
establish if there is an extensive As problem. It is far more
difficult to identify each affected well in view of the high
degree of spatial variability in arsenic concentrations
usually found in high-As areas.
Groundwater provides drinking water to more than
1.5 billion people daily and to many more in times of
surface water scarcity (DFID, 2001). While this review
has focussed on areas where groundwater As concentrations may be excessively high, it must not be forgotten
that this is the exception rather than the rule. Groundwater frequently provides a perfectly safe and reliable
form of drinking water. The challenge therefore is to
locate existing affected wells as quickly as possible, to
provide an alternative source of safe drinking water, and
to ensure that new sources are not As-rich. In most
aquifers, most wells are likely to be unaffected by As
enrichment, even when the groundwaters contain high
concentrations of dissolved Fe. Therefore it is also
necessary to understand why these groundwaters are not
affected. It appears that it is only when a number of
critical factors are combined that high-As groundwaters
are found.
While the authors have attempted to explain some of
the factors that give rise to high-As groundwaters, they
are aware that much remains unknown about exactly
how such waters are formed and that the generalisations
may not apply universally. They should serve as
hypotheses to be tested and amended by further detailed
field and laboratory investigations.
Acknowledgements
We thank Charles Harvey, Kirk Nordstrom, Don
Runnells, Alan Welch, Rick Johnston and an anonymous reviewer for constructive reviews, and the
Department for International Development (UK)
and the World Health Organization for support in
preparing this review. The review is published with the
permission of the Director, British Geological Survey
(NERC).
References
Abdullah, M.I., Shiyu, Z., Mosgren, K., 1995. Arsenic and
selenium species in the oxic and anoxic waters of the Oslofjord, Norway. Mar. Pollut. Bull. 31, 116–126.
Acharyya, S.K., 1997. Arsenic in groundwater—geological
overview. In: Consultation on Arsenic in Drinking Water
and Resulting Arsenic Toxicity in India and Bangladesh.
Proc. WHO conference, New Delhi, May 1997.
AGRG, 1978. The Wolfson Geochemical Atlas of England and
Wales. Clarendon Press, Oxford.
Aggett, J., Kriegman, M.R., 1988. The extent of formation of
arsenic(III) in sediment interstitial waters and its release to
hypolimnetic waters in Lake Ohakuri. Water Res. 22, 407–
411.
Aggett, J., O’Brien, G.A., 1985. Detailed model for the mobility of arsenic in lacustrine sediments based on measurements
in Lake Ohakuri. Environ. Sci. Technol. 19, 231–238.
Aggett, J., Roberts, L.S., 1986. Insight into the mechanism of
accumulation of arsenate and phosphate in hydro lake sediments by measuring the rate of dissolution with ethylenediaminetetraacetic acid. Environ. Sci. Technol. 20, 183–186.
Allan, R.J., Ball, A.J., 1990. An overview of toxic contaminants in water and sediments of the Great Lakes. Part I.
Water Pollut. Res. J. Can. 25, 387–505.
Amasa, S.K., 1975. Arsenic pollution at Obuasi goldmine,
town and surrounding countryside. Environ. Health Perspect. 12, 131–135.
Anderson, M.A., Ferguson, J.F., Gavis, J., 1976. Arsenate
adsorption on amorphous aluminum hydroxide. J. Colloid
Interface Sci. 54, 391–399.
Andreae, M.O., 1979. Arsenic speciation in seawater and
interstitial waters: the influence of biological-chemical interactions on the chemistry of a trace element. Limnol. Oceanog. 24, 440–452.
Andreae, M.O., 1980. Arsenic in rain and the atmospheric mass
balance of arsenic. J. Geophys. Res. 85, 4512–4518.
Andreae, M.O., Andreae, T.W., 1989. Dissolved arsenic species
in the Schelde estuary and watershed, Belgium. Estuar.
Coast. Shelf Sci. 29, 421–433.
Andreae, M.O., Byrd, T.J., Froelich, O.N., 1983. Arsenic,
antimony, germanium and tin in the Tejo estuary, Portugal:
modelling of a polluted estuary. Environ. Sci. Technol. 17,
731–737.
Appelo, C.A.J., Postma, D., 1994. Geochemistry, Groundwater
and Pollution. AA Balkema, Rotterdam.
Arehart, G.B., Chryssoulis, S.L., Kesler, S.E., 1993. Gold and
arsenic in iron sulfides from sediment-hosted disseminated
gold deposits-implications for depositional processes. Econ.
Geol. Bull. Soc. Econ. Geol. 88, 171–185.
Armienta, M.A., Rodriguez, R., Aguayo, A., Ceniceros, N.,
Villasenor, G., Cruz, O., 1997. Arsenic contamination of
groundwater at Zimapan, Mexico. Hydrogeol. J. 5, 39–46.
Arribére, M.A., Cohen, I.M., Ferpozzi, L.H., Kestelman, A.J.,
Casa, V.A., Guevara, S.R., 1997. Neutron activation analysis
of soils and loess deposits, for the investigation of the origin
of the natural arsenic-contamination in the Argentine
Pampa. Radiochim. Acta 78, 187–191.
Azcue, J.M., Nriagu, J.O., 1995. Impact of abandoned mine
tailings on the arsenic concentrations in Moira Lake,
Ontario. J. Geochem. Explor. 52, 81–89.
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
Azcue, J.M., Murdoch, A., Rosa, F., Hall, G.E.M., 1994.
Effects of abandoned gold mine tailings on the arsenic concentrations in water and sediments of Jack of Clubs Lake,
BC. Environ. Technol. 15, 669–678.
Azcue, J.M., Mudroch, A., Rosa, F., Hall, G.E.M., Jackson,
T.A., Reynoldson, T., 1995. Trace elements in water, sediments, porewater, and biota polluted by tailings from an
abandoned gold mine in British Columbia, Canada. J. Geochem. Explor. 52, 25–34.
Baes, C.F., Sharp, R.D., 1983. A proposal for estimation of soil
leaching and leaching constants for use in assessment models.
J. Environ. Qual. 12, 17–28.
Ball, J.W., Nordstrom, D.K., Jenne, E.A., Vivit, D.V.,
1998. Chemical Analyses Of Hot Springs, Pools, Geysers,
And Surface waters from Yellowstone National Park,
Wyoming, and Vicinity, 1974–1975. USGS Open-File Rep.
98–182.
Barbaris, B., Betterton, E.A., 1996. Initial snow chemistry survey of the Mogollon Rim in Arizona. Atmos. Environ. 30,
3093–3103.
Baur, W.H., Onishi, B.-M.H., 1969. Arsenic. In: Wedepohl,
K.H. (Ed.), Handbook of Geochemistry. Springer-Verlag,
Berlin. pp. 33-A-1–33-0-5.
Belkin, H.E., Zheng, B., Finkelman, R.B., 2000. Human health
effects of domestic combustion of coal in rural China: a
causal factor for arsenic and fluorine poisoning. In: 2nd
World Chinese Conf. Geological Sciences, Extended Abstr.,
August 2000, Stanford Univ., pp. 522–524.
Belzile, N., Tessier, A., 1990. Interactions between arsenic and
iron oxyhydroxides in lacustrine sediments. Geochim. Cosmochim. Acta 54, 103–109.
Benson, L.V., Spencer, R.J., 1983. A Hydrochemical Reconnaissance Study of the Walker River Basin, California and
Nevada. USGS Open File Rep., 83–740. United States
Geological Survey, Denver.
Berg, M., Tran, H.C., Nguyen, T.C., Pham, H.V., Schertenleib,
R., Giger, W., 2001. Arsenic contamination of groundwater
and drinking water in Vietnam: a human health threat.
Environ. Sci. Technol. 35, 2621–2626.
Berner, R., 1981. A new geochemical classification of sedimentary environments. J. Sed. Petrol. 51, 359–365.
BGS, DPHE, 2001. Arsenic contamination of groundwater in
Bangladesh. In: Kinniburgh, D.G., Smedley, P.L. (Eds.),
British Geological Survey (Technical Report, WC/00/19. 4
Volumes). British Geological Survey, Keyworth.
Boughriet, A., Figueiredo, R.S., Laureyns, J., Recourt, P.,
1997. Identification of newly generated iron phases in recent
anoxic sediments: Fe-57 Mössbauer and microRaman spectroscopic studies. Journal of the Chem. Soc.-Faraday Trans.
93, 3209–3215.
Bowell, R.J., 1992. Supergene gold mineralogy at Ashanti,
Ghana: implications for the supergene behaviour of gold.
Mineral. Mag. 56, 545–560.
Bowell, R.J., 1993. Mineralogy and geochemistry of tropical
rain forest soils: Ashanti, Ghana. Chem. Geol. 106, 345–348.
Bowell, R.J., 1994. Sorption of arsenic by iron-oxides and
oxyhydroxides in soils. Appl. Geochem. 9, 279–286.
Boyle, D.R., Turner, R.J.W., Hall, G.E.M., 1998. Anomalous
arsenic concentrations in groundwaters of an island community, Bowen Island, British Columbia. Environ. Geochem.
Health 20, 199–212.
561
Boyle, R.W., Jonasson, I.R., 1973. The geochemistry of As and
its use as an indicator element in geochemical prospecting. J.
Geochem. Explor. 2, 251–296.
Brannon, J.M., Patrick, W.H., 1987. Fixation, transformation,
and mobilization of arsenic in sediments. Environ. Sci.
Technol. 21, 450–459.
Bright, D.A., Dodd, M., Reimer, K.J., 1996. Arsenic in subArctic lakes influenced by gold mine effluent: the occurrence
of organoarsenicals and ‘hidden’ arsenic. Sci. Tot. Environ.
180, 165–182.
Brookins, D.G., 1988. Eh-pH Diagrams for Geochemistry.
Springer-Verlag, Berlin.
Cáceres, L., Gruttner, E., Contreras, R., 1992. Water recycling
in arid regions—Chilean case. Ambio 21, 138–144.
Canfield, D.E., 1989. Reactive iron in sediments. Geochim.
Cosmochim. Acta 53, 619–632.
Canfield, D.E., Berner, R.A., 1987. Dissolution and pyritization of magnetite in anoxic marine sediments. Geochim.
Cosmochim. Acta 51, 645–659.
Cebrián, M.E., Albores, M.A., Garciá-Vargas, G., Del Razo,
L.M., Ostrosky-Wegman, P., 1994. Chronic arsenic poisoning
in humans. In: Nriagu, J.O. (Ed.), Arsenic in the Environment, Part II: Human Health and Ecosystem Effects. John
Wiley, New York, pp. 93–107.
CGWB, 1999. High Incidence of Arsenic in Groundwater in
West Bengal. Central Ground Water Board, India, Ministry
of Water Resources, Government of India.
Chen, C.J., Chuang, Y.C., Lin, T.M., Wu, H.Y., 1985. Malignant neoplasms among residents of a blackfoot diseaseendemic area in Taiwan: high-arsenic artesian well water and
cancers. Cancer Res. 45, 5895–5899.
Chen, S.L., Dzeng, S.R., Yang, M.H., Chlu, K.H., Shieh,
G.M., Wal, C.M., 1994. Arsenic species in groundwaters of
the Blackfoot disease areas, Taiwan. Environ. Sci.Technol.
28, 877–881.
Chen, S.L., Yeh, S.J., Yang, M.H., Lin, T.H., 1995. Trace element concentration and arsenic speciation in the well water
of a Taiwan area with endemic Blackfoot disease. Biol. Trace
Elem. Res. 48, 263–274.
Cherry, J.A., Shaikh, A.U., Tallman, D.E., Nicholson, R.V.,
1979. Arsenic species as an indicator of redox conditions in
groundwater. J. Hydrol. 43, 373–392.
Choprapawon, C., Rodcline, A., 1997. Chronic arsenic poisoning in Ronpibool Nakhon Sri Thammarat, the Southern
Province of Thailand. In: Abernathy, C.O., Calderon, R.L.,
Chappell, W.R. (Eds.), Arsenic Exposure and Health Effects.
Chapman and Hall, London, pp. 69–77.
Collon, P., Kutschera, W., Loosli, H.H., Lehmann, B.E.,
Purtschert, R., Love, A., Sampson, L., Anthony, D., Cole,
D., Davids, B., Morrissey, D.J., Sherrill, B.M., Steiner, M.,
Pardo, R.C., Paul, M., 2000. Kr-81 in the Great Artesian
Basin, Australia: a new method for dating very old groundwater. Earth Planet. Sci. Lett. 182, 103–113.
Cook, S.J., Levson, V.M., Giles, T.R., Jackaman, W., 1995.
A comparison of regional lake sediment and till geochemistry surveys—a case-study from the Fawnie Creek
area, Central British Columbia. Explor. Min. Geol. 4, 93–
110.
Cornell, R.M., Schwertmann, U., 1996. The Iron Oxides:
Structure, Properties, Reactions, Occurrence and Uses.
VCH, Weinheim.
562
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
Crecelius, E.A., 1975. The geochemical cycle of arsenic in Lake
Washington and its relation to other elements. Limnol.
Oceanog. 20, 441–451.
Criaud, A., Fouillac, C., 1989. The distribution of arsenic(III)
and arsenic(V) in geothermal waters: Examples from the
Massif Central of France, the Island of Dominica in the
Leeward Islands of the Caribbean, the Valles Caldera of New
Mexico, USA, and southwest Bulgaria. Chem. Geol. 76,
259–269.
Cronan, D.S., 1972. The mid-Atlantic Ridge near 45 N, XVII:
Al, As, Hg, and Mn in ferriginous sediments from the median valley. Can. J. Earth Sci. 9, 319–323.
CSME, 1997. Geology And Geochemistry Of Arsenic Occurrences In Groundwater Of Six Districts of West Bengal.
Report of the Centre for Study of Man and Environment,
Calcutta.
Cullen, W.R., Reimer, K.J., 1989. Arsenic speciation in the
environment. Chem. Rev. 89, 713–764.
Cummings, D.E., Caccavo, F., Fendorf, S., Rosenzweig, R.F.,
1999. Arsenic mobilization by the dissimilatory Fe(III)reducing bacterium Shewanella alga BrY. Environ. Sci.
Technol. 33, 723–729.
Cutter, G.A., Cutter, L.S., Featherstone, A.M., Lohrenz, S.E.,
2001. Antimony and arsenic biogeochemistry in the western
Atlantic Ocean. Deep-Sea Res. Part II—Topical Stud. Oceanog. 48, 2895–2915.
Das, D., Chatterjee, A., Mandal, B.K., Samanta, G., Chakraborti, D., Chanda, B., 1995. Arsenic in ground-water in 6
districts of West Bengal, India—the biggest arsenic calamity
in the world. 2. Arsenic concentration in drinking-water,
hair, nails, urine, skin-scale and liver-tissue, biopsy of the
affected people. Analyst 120, 917–924.
Datta, D.K., Subramanian, V., 1997. Texture and mineralogy
of sediments from the Ganges-Brahmaputra-Meghna river
system in the Bengal basin, Bangladesh and their environmental implications. Environ. Geol. 30, 181–188.
Davis, A., de Curnou, P., Eary, L.E., 1997. Discriminating
between sources of arsenic in the sediments of a tidal waterway,
Tacoma, Washington. Environ. Sci. Technol. 31, 1985–1991.
Davison, W., 1993. Iron and manganese in lakes. Earth Sci.
Rev. 34, 119–163.
Del Razo, L.M., Arellano, M.A., Cebrián, M.E., 1990. The
oxidation states of arsenic in well-water from a chronic
arsenicism area of northern Mexico. Environ. Pollut. 64,
143–153.
Del Razo, L.M., Hernández, J.L., Garcı́a-Vargas, G.G.,
Ostrosky-Wegman, P., Cortinas de Nava, C., Cebrián, M.E.,
1994. Urinary excretion of arsenic species in a human population chronically exposed to arsenic via drinking water. A
pilot study. In: Chappell, W.R., Abernathy, C.O., Cothern,
C.R. (Eds.), Arsenic Exposure and Health, 91–100. Science
and Technology Letters, Northwood.
Deuel, L.E., Swoboda, A.R., 1972. Arsenic solubility in a
reduced environment. Soil Sci. Soc. Am. Proc. 36, 276–278.
De Vitre, R., Belzile, N., Tessier, A., 1991. Speciation and
adsorption of arsenic on diagenetic iron oxyhydroxides.
Limnol. Oceanog. 36, 1480–1485.
DFID, 2001. Addressing the Water Crisis: Healthier And More
Productive Lives For People. UK Department for International Development, London.
(special issue). Curr. Sci. 734859.
DPHE/BGS/MML, 1999. Groundwater Studies for Arsenic
Contamination in Bangladesh. Phase I: Rapid Investigation
Phase. BGS/MML Technical Report to Department for
International Development, UK), 6 volumes.
Driehaus, W., Jekel, M., Hildebrandt, U., 1998. Granular
ferric hydroxide-a new adsorbent for the removal of arsenic
from natural water. J. Water Supp. Res. Technol.-Aqua 47,
30–35.
Driehaus, W., Seith, R., Jekel, M., 1995. Oxidation of arsenate(III) with manganese oxides in water-treatment. Water Res.
29, 297–305.
Drodt, M., Trautwein, A.X., Konig, I., Suess, E., Koch, C.B.,
1997. Mössbauer spectroscopic studies on the iron forms of
deep-sea sediments. Physics and Chemistry of Minerals 24,
281–293.
Dudas, M.J., 1984. Enriched levels of arsenic in post-active acid
sulfate soils in Alberta. Soil Sci. Soc. Am. J. 48, 1451–1452.
Dudas, M.J., Warren, C.J., Spiers, G.A., 1988. Chemistry of
arsenic in acid sulfate soils of northern Alberta. Comm. Soil
Sci. Plant Anal. 19, 887–895.
Durum, W.H., Hem, J.D., Heidel, S.G., 1971. Reconnaissance
of Selected Minor Elements in Surface Waters of the United
States, October 1970. US Geol. Surv. Circ. 643.
Dzombak, D.A., Morel, F.M.M., 1990. Surface Complexation
Modelling- Hydrous Ferric Oxide. John Wiley, New York.
Eary, L.E., Schramke, J.A., 1990. Rates of inorganic oxidation
reactions involving dissolved oxygen. In: Melchior, D.C.,
Bassett, R.L. (Eds.), Chemical Modeling of Aqueous Systems II. ACS Symposium Series. American Chemical
Society, Washington, DC, pp. 379–396.
Edmunds, W.M., Cook, J.M., Kinniburgh, D.G., Miles, D.L.,
Trafford, J.M., 1989. Trace-element Occurrence in British
Groundwaters. Res. Report SD/89/3, British Geological
Survey, Keyworth.
Edwards, M., 1994. Chemistry of arsenic removal during coagulation and Fe–Mn oxidation. J. Am. Water Works Assoc.
86, 64–78.
Edwards, M., Patel, S., McNeill, L., Chen, H.W., Frey, M.,
Eaton, A.D., Antweiler, R.C., Taylor, H.E., 1998. Considerations in As analysis and speciation. J. Am. Water
Works Assoc. 90, 103–113.
Ellis, A.J., Mahon, W.A.J., 1977. Chemistry and Geothermal
Systems. Academic Press, New York.
Erhlich, H.L., 1996. Geomicrobiology, 3rd ed. Marcel Dekker,
New York.
Farmer, J.G., Baileywatts, A.E., Kirika, A., Scott, C., 1994.
Phosphorus fractionation and mobility in Loch Leven sediments. Aquat. Conserv.-Mar. Freshwater Ecosys. 4, 45–56.
Fendorf, S., Eick, M.J., Grossel, P., Sparks, D.L., 1997. Arsenate and chromate retention mechanisms on goethite. 1.
Surface structure. Environ. Sci. Technol. 31, 315–320.
Ficklin, W. H. and Callender, E., 1989. Arsenic geochemistry
of rapidly accumulating sediments, Lake Oahe, South
Dakota. In: Mallard, G. E. and Ragone, S. E. (Eds.), US
Geol. Surv. Water Resour. Investig. Rep. 88–4420. U.S.
Geol. Surv. Toxic Substances Hydrology Program—Proc.
Tech. Meeting, Phoenix, Arizona, 26–30 September 1988, pp.
217–222.
Finkelman, R.B., Belkin, H.E., Zheng, B., 1999. Health
impacts of domestic coal use in China. Proc. Nat. Acad. Sci.,
USA 96, 3427–3431.
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
Fleet, M.E., Mumin, A.H., 1997. Gold-bearing arsenian pyrite
and marcasite and arsenopyrite from Carlin Trend gold
deposits and laboratory synthesis. Am. Mineral. 82, 182–193.
Fontaine, J.A., 1994. Chapter 28. Regulating arsenic in Nevada
drinking water supplies: past problems, future challenges. In:
Chappell, W.R., Abernathy, C.O., Cothern, C.R. (Eds.),
Arsenic Exposure and Health. Science and Technology Letters,
Northwood, pp. 285–288.
Forstner, U., Haase, I., 1998. Geochemical demobilization of
metallic pollutants in solid waste-implications for arsenic in
waterworks sludges. J. Geochem. Explor. 62, 29–36.
Foster, A.L., Breit, G.N., Welch, A.L., Whitney, J.W., Islam,
M. Nazrul, Islam, H. Khairul, Alam M.K., Islam, M.S.,
2000. In-situ Identification of Arsenic Species in Soil and
Aquifer Sediment from Ramrail, Brahmanbaria, Bangladesh.
Presentation at AGU Fall meeting, 15–19 December 2000,
San Francisco, California.
Foster, A.L., Brown, G.E., Parks, G.A., 1998. X-ray absorption fine-structure spectroscopy study of photocatalyzed,
heterogeneous As(III) oxidation on kaolin and anatase.
Environ. Sci. Technol 32, 1444–1452.
Fredrickson, J.K., Zachara, J.M., Kennedy, D.W., Dong, H.,
Onstott, T.C., Hinman, N.W., Li, S.-M., 1998. Biogenic iron
mineralization accompanying the dissimilatory reduction of
hydrous ferric oxide by a groundwater bacterium. Geochim.
Cosmochim. Acta 62, 3239–3257.
Froelich, P.N., Kaul, L.W., Byrd, J.T., Andreae, M.O., Roe,
K.K., 1985. Arsenic, barium, germanium, tin, dimethyl-sulfide and nutrient biogeochemsitry in Charlotte Harbour,
Florida, a phosphorus-enriched estuary. Estuar. Coast. Shelf
Sci. 20, 239–264.
Fujii, R., Swain, W.C., 1995. Areal Distribution of Selected Trace
Elements, Salinity, and Major Ions in Shallow Ground Water,
Tulare Basin, Southern San Joaquin Valley, California. US
Geol. Surv. Water-Resour. Investig. Rep. 95–4048.
Gelova, G.A., 1977. Hydrogeochemistry of Ore Elements.
Nedra, Moscow.
Génin, J.-M.R., Bourrié, G., Trolard, F., Abdelmoula, M.,
Jaffrezic, A., Refait, P., Maitre, V., Humbert, B., Herbillon,
A., 1998. Thermodynamic equilibria in aqueous suspensions
of synthetic and natural Fe(II)-Fe(III) green rusts: occurrences of the mineral in hydromorphic soils. Environ. Sci.
Technol. 32, 1058–1068.
Ghosh, M.M., Yuan, J.R., 1987. Adsorption of inorganic
arsenic and organoarsenicals on hydrous oxides. Environ.
Prog. 6, 150–157.
Goldberg, S., 1986. Chemical modeling of arsenate adsorption
on aluminum and iron oxide minerals. Soil Sci. Soc. Am. J.
50, 1154–1157.
Goldberg, S., Glaubig, R.A., 1988. Anion sorption on a calcareous, montmorillonitic soil—arsenic. Soil Sci. Soc. Am. J.
52, 1297–1300.
Gómez-Ariza, J.L., Sanchez Rodas, D., Giraldez, I., 1998.
Selective extraction of iron oxide associated arsenic species
from sediments for speciation with coupled HPLC-HG-AAS.
Journal of Anal. Atom. Spectrom. 13, 1375–1379.
Grimes, D.J., Ficklin, W.H., Meier, A.L., McHugh, J.B., 1995.
Anomalous gold, antimony, arsenic, and tungsten in ground
water and alluvium around disseminated gold deposits along
the Getchell Trend, Humboldt County, Nevada. J. Geochem.
Explor. 52, 351–371.
563
Guerin, W.F., Blakemore, R.P., 1992. Redox cycling of iron
supports growth and magnetite synthesis by Aquaspirillum
magnetotacticum. Appl. Environ. Microbiol. 58, 1102–
1109.
Gulens, J., Champ, D.R., Jackson, R.E., 1979. Influence of
redox environments on the mobility of arsenic in ground
water. In: Jenne, E.A. (Ed.), Chemical Modelling in Aqueous
Systems. American Chemical Society, pp. 81–95.
Guo, H.R., Chen, C.J., Greene, H.L., 1994. Arsenic in drinking
water and cancers: a brief descriptive view of Taiwan studies.
In: Chappell, W.R., Abernathy, C.O., Cothern, C.R. (Eds.),
Arsenic Exposure and Health. Science and Technology Letters, Northwood, pp. 129–138.
Guo, T.Z., DeLaune, R.D., Patrick, W.H., 1997. The influence
of sediment redox chemistry on chemically active forms of
arsenic, cadmium, chromium, and zinc in estuarine sediment.
Environ. Internat 23, 305–316.
Gurzau, E.S., Gurzau, A.E., 2001. Arsenic in drinking water
from groundwater in Transylvania, Romania: In: Chapell,
W.R., Abernathy, C.O., Calderon, R.L. (Eds.), Arsenic
Exposure and Health Effects IV. Elsevier, Amsterdam, pp.
181–184.
Gustafsson, J.P., Jacks, G., 1995. Arsenic geochemistry in
forested soil profiles as revealed by solid-phase studies. Appl.
Geochem. 10, 307–315.
Gustafsson, J.P., Tin, N.T., 1994. Arsenic and selenium in
some Vietnamese acid sulfate soils. Sci. Tot. Environ. 151,
153–158.
Hale, J.R., Foos, A., Zubrow, J.S., Cook, J., 1997. Better
characterization of arsenic and chromium in soils: a fieldscale example. J. Soil Contam. 6, 371–389.
Hall, G.E.M., Pelchat, J.C., Gauthier, G., 1999. Stability of
inorganic arsenic(III) and arsenic(V) in water samples. J.
Anal. Atom. Spectrom 14, 205–213.
Hasegawa, H., 1997. The behavior of trivalent and pentavalent
methylarsenicals in Lake Biwa. Appl. Organometal. Chem.
11, 305–311.
Hasegawa, H., Matsui, M., Okamura, S., Hojo, M., Iwasaki,
N., Sohrin, Y., 1999. Arsenic speciation including ‘hidden’
arsenic in natural waters. Applied Organometal. Chem. 13,
113–119.
Hess, R.E., Blanchar, R.W., 1977. Dissolution of arsenic from
waterlogged and aerated soil. Soil Sci. Soc. Am. J. 41, 861–
865.
Heinrichs, G., Udluft, P., 1999. Natural arsenic in Triassic
rocks: a source of drinking-water contamination in Bavaria,
Germany. Hydrogeol. J. 7, 468–476.
Hiemstra, T., van Riemsdijk, W.H., 1996. A surface structural
approach to ion adsorption: the charge distribution, CD.
Model. J. Colloid Interface Sci. 179, 488–508.
Hiemstra, T., van Riemsdijk, W.H., 1999. Surface structural
ion adsorption modeling of competitive binding of oxyanions
by metal hydroxides. J. Colloid Interface Sci. 210, 182–193.
Hingston, F.J., Posner, A.M., Quirk, J.P., 1971. Competitive
adsorption of negatively charged ligands on oxide surfaces.
In: Discussions of the Faraday Society, No. 52. The Faraday
Society, London, pp. 334–342.
Hopenhayn-Rich, C., Biggs, M.L., Fuchs, A., Bergoglio, R.,
Tello, E.E., Nicolli, H., Smith, A.H., 1996. Bladder-cancer
mortality associated with arsenic in drinking water in
Argentina. Epidemiol. 7, 117–124.
564
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
Howard, A.G., Apte, S.C., Comber, S.D.W., Morris, R.J.,
1988. Biogeochemical control of the summer distribution and
speciation of arsenic in the Tamar estuary. Estuar. Coast.
Shelf Sci. 27, 427–443.
Howard, A.G., Hunt, L.E., Salou, C., 1999. Evidence supporting the presence of dissolved dimethylarsinate in the marine
environment. Appl. Organometal. Chem. 13, 39–46.
Hsu, K.-H., Froines, J.R., Chen, C.-J., 1997. Studies of arsenic
ingestion from drinking water in northeastern Taiwan: chemical speciation and urinary metabolites. In: Abernathy, C.O.,
Calderon, R.L., Chappell, W.R. (Eds.), Arsenic Exposure and
Health Effects. Chapman Hall, London, pp. 190–209.
Jain, A., Raven, K.P., Loeppert, R.H., 1999. Arsenite and
arsenate adsorption on ferrihydrite: surface charge reduction
and net OH- release stoichiometry. Environ. Sci. Technol. 33,
1179–1184.
James, H.L., 1966. Chemistry of the iron-rich sedimentary
rocks. Data of Geochemistry. USGS Prof. Pap. 440-W.
Johnson, D.L., Pilson, M.E.Q., 1975. The oxidation of arsenite
in seawater. Environ. Lett. 8, 157–171.
Jonnalagadda, S.B., Nenzou, G., 1996. Studies on arsenic-rich
mine dumps. 1. Effect on the surface soil. J. Environ. Sci. Health
Part A—Environ. Sci. Toxic Hazard. Subst. Control 31, 1909–
1915.
Karcher, S., Cáceres, L., Jekel, M., Contreras, R., 1999.
Arsenic removal from water supplies in Northern Chile using
ferric chloride coagulation. J. Chart. Inst. Water Environ.
Manag. 13, 164–169.
Kavanagh, P.J., Farago, M.E., Thornton, I., Braman, R.S.,
1997. Bioavailability of arsenic in soil and mine wastes of the
Tamar Valley, SW England. Chem. Spec. Bioavail. 9, 77–81.
Kent, D.B., Davis, J.A., Anderson, L.C.D., Rea, B.A., 1995.
Transport of chromium and selenium in a pristine sand and
gravel aquifer: role of adsorption processes. Water Resour.
Res. 31, 1041–1050.
Keon, N.E., Swartz, C.H., Brabander, D.J., Harvey, C.,
Hemond, H.F., 2001. Validation of an arsenic sequential
extraction method for evaluating mobility in sediments.
Environ. Sci. Technol. 35, 2778–2784.
Khalid, R.A., Patrick, W.H., DeLaune, R.D., 1977. Phosphorus sorption characteristics of flooded soils. Soil Sci. Soc.
Am. 41, 305–310.
Kinniburgh, D.G., Sridhar, K., Jackson, M.L., 1977. Specific
adsorption of zinc and cadmium by iron and aluminum hydrous oxides. In: Hanford, W.A. (Ed.), The Fifteenth Hanford
Life Sciences Symposium on Biological Implications of Metals
in the Environment, 29 September–1 October 1975. USERDA,
Natl Technical Information Service, 1977, Springfield, VA.
Klumpp, D.W., Peterson, P.J., 1979. Arsenic and other trace
elements in the waters and organisms of an estuary in SW
England. Environ. Pollut. 19, 11–20.
Kneebone, P.E., Hering, J.G., 2000. Behavior of arsenic and
other redox sensitive elements in Crowley Lake, CA: a
reservoir in the Los Angeles Aqueduct system. Environ. Sci.
Technol. 34, 4307–4312.
Komnitsas, K., Xenidis, A., Adam, K., 1995. Oxidation of
pyrite and arsenopyrite in sulphidic spoils in Lavrion. Miner.
Engineer 12, 1443–1454.
Korte, N., 1991. Naturally occurring arsenic in groundwaters
of the midwestern United States. Environ. Geol. Water Sci.
18, 137–141.
Korte, N.E., Fernando, Q., 1991. A review of arsenic(III) in
groundwater. Crit. Rev. Environ. Control 21, 1–39.
Krause, E., Ettel, V.A., 1989. Solubilities and stabilities of ferric arsenate compounds. Hydrometal. 22, 311–337.
Kuhlmeier, P.D., 1997a. Partitioning of arsenic species in finegrained soils. J. Air Waste Manag. Assoc. 47, 481–490.
Kuhlmeier, P.D., 1997b. Sorption and desorption of arsenic
from sandy soils: Column studies. J. Soil Contam. 6, 21–
36.
Kuhn, A., Sigg, L., 1993. Arsenic cycling in eutrophic Lake
Greifen, Switzerland—influence of seasonal redox processes.
Limnol. Oceanog. 38, 1052–1059.
Kuo, T.-L., 1968. Arsenic content of artesian well water in
endemic area of chronic arsenic poisoning. Rep. Inst. Pathol.
National Taiwan Univ. 20, 7–13.
Langmuir, D., 1997. Aqueous Environmental Geochemistry.
Prentice Hall, New Jersey.
Legeleux, F., Reyss, J.L., Bonte, P., Organo, C., 1994. Concomitant enrichments of uranium, molybdenum and arsenic
in suboxic continental-margin sediments. Oceanolog. Acta
17, 417–429.
Lenvik, K., Steinnes, E., Pappas, A.C., 1978. Contents of some
heavy metals in Norwegian rivers. Nord. Hydrol. 9, 197–206.
Lerda, D.E., Prosperi, C.H., 1996. Water mutagenicity and
toxicology in Rio Tercero, Cordoba, Argentina. Water Res.
30, 819–824.
Lindberg, R.D., Runnells, D.D., 1984. Ground water redox
reactions: an analysis of equilibrium state applied to Eh
measurements and geochemical modeling. Sci. 225, 925–
927.
Livesey, N.T., Huang, P.M., 1981. Adsorption of arsenate by
soils and its relation to selected chemical properties and
anions. Soil Sci. 131, 88–94.
Lo, M.-C., Hsen, Y.-C., Lin, K.-K., 1977. Second Report on
the Investigation of Arsenic Content in Underground Water
in Taiwan. Taiwan Provincial Institute of Environmental
Sanitation, Taichung, Taiwan.
Lovley, D.R., Chapelle, F.H., 1995. Deep subsurface microbial
processes. Rev. Geophys. 33, 365–381.
Luo, Z.D., Zhang, Y.M., Ma, L., Zhang, G.Y., He, X., Wilson,
R., Byrd, D.M., Griffiths, J.G., Lai, S., He, L., Grumski, K.,
Lamm, S.H., 1997. Chronic arsenicism and cancer in Inner
Mongolia - consequences of well-water arsenic levels greater
than 50 mg l 1. In: Abernathy, C.O., Calderon, R.L., Chappell, W.R. (Eds.), Arsenic Exposure and Health Effects.
Chapman Hall, London, pp. 55–68.
Ma, H.Z., Xia, Y.J., Wu, K.G., Sun, T.Z., Mumford, J.L.,
1999. Arsenic exposure and health effects in Bayingnormen,
Inner Mongolia. In: Chappell, W.R., Abernathy, C.O., Calderon, R.L. (Eds.), Arsenic Exposure and Health Effects.
Elsevier, Amsterdam, pp. 127–131.
Maest, A.S., Pasilis, S.P., Miller, L.G., Nordstrom, D.K., 1992.
Redox geochemistry of arsenic and iron in Mono Lake,
California, USA. In: Kharaka, Y.K., Maest, A.S. (Eds.),
Proc. 7th Internat. Symp. Water-Rock Interaction. A. A.
Balkema, Rotterdam, pp. 507–511.
Maher, B.A., Taylor, R.M., 1988. Formation of ultrafinegrained magnetite in soils. Nature 336, 368–370.
Maher, W.A., 1985. Arsenic in coastal waters of South Australia. Water Res. 19, 933–934.
Manning, B.A., Fendorf, S.E., Goldberg, S., 1998. Surface
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
structures and stability of arsenic(III) on goethite: spectroscopic evidence for inner-sphere complexes. Environ. Sci.
Technol. 32, 2383–2388.
Manning, B.A., Goldberg, S., 1996. Modeling competitive
adsorption of arsenate with phosphate and molybdate on
oxide minerals. Soil Sci. Soc. Am. J. 60, 121–131.
Manning, B.A., Goldberg, S., 1997a. Arsenic(III) and arsenic(V)
adsorption on three California soils. Soil Sci. 162, 886–895.
Manning, B.A., Goldberg, S., 1997b. Adsorption and stability
of arsenic(III) at the clay mineral-water interface. Environ.
Sci. Technol. 31, 2005–2011.
Martin, J.-M., Whitfield, M., 1983. The significance of the river
input of chemical elements to the ocean. In: Wong, C.S., Boyle,
E., Bruland, K.W., Burton, J.D., Goldberg, E.D. (Eds.), Trace
Metals in Seawater. Plenum Press, New York, pp. 265–296.
Masscheleyn, P.H., DeLaune, R.D., Patrick, W.H., 1991. Effect of
redox potential and pH on arsenic speciation and solubility in a
contaminated soil. Environ. Sci. Technol. 25, 1414–1419.
Matis, K.A., Zouboulis, A.I., Malamas, F.B., Afonso, M.D.R.,
Hudson, M.J., 1997. Flotation removal of As(V) onto goethite. Environ. Pollut. 97, 239–245.
Matis, K.A., Zouboulis, A.I., Zamboulis, D., Valtadorou,
A.V., 1999. Sorption of As(V) by goethite particles and study
of their flocculation. Water Air Soil Pollut. 111, 297–316.
Matisoff, G., Khourey, C.J., Hall, J.F., Varnes, A.W., Strain,
W.H., 1982. The nature and source of arsenic in northeastern
Ohio groundwater. Ground Water 20, 446–456.
McCreadie, H., Blowes, D.W., Ptacek, C.J., Jambor, J.L.,
2000. Influence of reduction reactions and solid-phase composition on porewater concentrations of arsenic. Environ.
Sci. Technol. 34, 3159–3166.
McGeehan, S.L., 1996. Arsenic sorption and redox reactions:
Relevance to transport and remediation. J. Environ. Sci.
Health Part A-Environ. Sci. Engineer. Toxic Hazard. Subst.
Control 31, 2319–2336.
McGeehan, S.L., Naylor, D.V., 1994. Sorption and redox
transformation of arsenite and arsenate in two flooded soils.
Soil Sci. Soc. Am. J. 58, 337–342.
McLaren, S.J., Kim, N.D., 1995. Evidence for a seasonal fluctuation of arsenic in New Zealand’s longest river and the
effect of treatment on concentrations in drinking water.
Environ. Pollut. 90, 67–73.
Millero, F., Huang, F., Zhu, X.R., Liu, X.W., Zhang, J.Z.,
2001. Adsorption and desorption of phosphate on calcite and
aragonite in seawater. Aquat. Geochem. 7, 33–56.
Mok, W., Wai, C.M., 1990. Distribution and mobilization of
arsenic and antimony species in the Coeur D’Alene River,
Idaho. Environ. Sci. Technol. 24, 102–108.
Moore, J.N., Ficklin, W.H., Johns, C., 1988. Partitioning of
arsenic and metals in reducing sulfidic sediments. Environ.
Sci. Technol. 22, 432–437.
Mortimer, C.H., 1942. The exchange of dissolved substances
between mud and water in lakes. Parts III and IV, summary
and references. J. Ecol. 30, 147–201.
Nag, J.K., Balaram, V., Rubio, R., Alberti, J., Das, A.K., 1996.
Inorganic arsenic species in groundwater: a case study from
Purbasthali, Burdwan, India. J. Trace Elem. Med. Biol. 10,
20–24.
Nagorski, S.A., Moore, J.N., 1999. Arsenic mobilization in the
hyporheic zone of a contaminated stream. Water Resourc.
Res. 35, 3441–3450.
565
Navarro, M., Sanchez, M., Lopez, H., Lopez, M.C., 1993.
Arsenic contamination levels in waters, soils, and sludges in
southeast spain. Bull. Environ. Contam. Toxicol. 50, 356–
362.
Newman, D.K., Ahmann, D., Morel, F.M.M., 1998. A brief
review of microbial arsenate respiration. Geomicrobiol. 15,
255–268.
Nicolli, H.B., Merino, M.H., 2002. High contents of F, As and
Se in groundwater of the Carcarañá river basin, Argentine
Pampean Plain. Environ. Geol., in press.
Nicolli, H.B., Suriano, J.M., Peral, M.A.G., Ferpozzi, L.H.,
Baleani, O.A., 1989. Groundwater contamination with
arsenic and other trace-elements in an area of the Pampa,
province of Córdoba, Argentina. Environ. Geol. Water Sci.
14, 3–16.
Nicolli, H.B., Tineo, A., Garcı́a, J.W., Falcón, C.M., Merino,
M.H., 2001. Trace-element quality problems in groundwater
from Tucumán, Argentina. In: Cidu, R. (Ed.), Water-Rock
Interaction 2001. Vol. 2. Swets & Zeitlinger, Lisse, pp. 993–
996.
Nimick, D.A., Moore, J.N., Dalby, C.E., Savka, M.W., 1998.
The fate of geothermal arsenic in the Madison and Missouri
Rivers, Montana and Wyoming. Water Resour. Res. 34,
3051–3067.
Niu, S., Cao, S., Shen, E., 1997. The status of arsenic poisoning
in China. In: Abernathy, C.O., Calderon, R.L., Chappell,
W.R. (Eds.), Arsenic Exposure and Health Effects. Chapman
Hall, London, pp. 78–83.
Nordstrom, D.K., 2000. An overview of arsenic mass poisoning
in Bangladesh and West Bengal, India. In: Young, C. (Ed.),
Minor Elements 2000: Processing and Environmental
Aspects of As, Sb, Se, Te, and Bi. Society for Mining,
Metallurgy and Exploration, pp. 21–30.
Nordstrom, D.K., Alpers, C.N., 1999. Negative pH, efflorescent mineralogy, and consequences for environmental
restoration at the Iron Mountain Superfund Site, California.
Proc. Nat. Acad. Sci., USA 96, 3455–3462.
Nordstrom, D.K., Alpers, C.N., Ptacek, C.J., Blowes, D.W.,
2000. Negative pH and extremely acidic mine waters from
Iron Mountain California. Environ. Sci. Technol. 34, 254–
258.
Nriagu, J.O., Pacyna, J.M., 1988. Quantitative assessment of
worldwide contamination of air, water, and soils by trace
metals. Nature 333, 134–139.
Onishi, H., Sandell, E.B., 1955. Geochemistry of arsenic. Geochim. Cosmochim. Acta 7, 1–33.
Oremland, R.S., Dowdle, P.R., Hoeft, S., Sharp, J.O., Schaefer, J.F., Miller, L.G., Blum, J.S., Smith, R.L., Bloom, N.S.,
Wallschlaeger, D., 2000. Bacterial dissimilatory reduction of
arsenate and sulfate in meromictic Mono Lake, California.
Geochim. Cosmochim. Acta 64, 3073–3084.
Oscarson, D.W., Huang, P.M., Defosse, D., Herbillion, A.,
1981. Oxidative power of Mn(IV) and Fe(III) oxides with
respect to As(III) in terrestrial and aquatic environments.
Nature 291, 50–51.
Oscarson, D.W., Huang, P.M., Liaw, W.K., Hammer, U.T.,
1983. Kinetics of oxidation of arsenite by various manganese
dioxides. Soil Sci. Soc. Am. J. 47, 644–648.
Paige, C.R., Snodgrass, W.J., Nicholson, R.V., Scharer, J.M.,
1997. An arsenate effect on ferrihydrite dissolution kinetics
under acidic oxic conditions. Water Res. 31, 2370–2382.
566
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
Palmer, C.A., Klizas, S.A., 1997. The Chemical Analysis Of
Argonne Premium Coal Samples. US Geol. Surv. Bull. 2144,
US Geological Survey, Reston, USA.
Peterson, M.L., Carpenter, R., 1983. Biogeochemical processes
affecting total arsenic and arsenic species distributions in an
intermittently anoxic Fjord. Mar. Chem. 12, 295–321.
Peterson, M.L., Carpenter, R., 1986. Arsenic distributions in
porewaters and sediments of Puget Sound, Lake Washington, the Washington coast and Saanich Inlet, B.C. Geochim.
Cosmochim. Acta 50, 353–369.
Pettine, M., Camusso, M., Martinotti, W., 1992. Dissolved and
particulate transport of arsenic and chromium in the Po
River, Italy. Sc. Tot. Environ. 119, 253–280.
PHED, 1991. Arsenic Pollution in Groundwater in West Bengal. Report of Arsenic Investigation Project to the National
Drinking Water Mission, Delhi, India.
PHED/UNICEF, 1999. Joint Plan of Action to Address
Arsenic Contamination of Drinking Water. Government of
West Bengal and UNICEF. Public Health Engineering
Department, Government of West Bengal.
Pichler, T., Veizer, J., Hall, G.E.M., 1999. Natural input of
arsenic into a coral reef ecosystem by hydrothermal fluids
and its removal by Fe(III) oxyhydroxides. Environ. Sci.
Technol. 33, 1373–1378.
Pierce, M.L., Moore, C.B., 1982. Adsorption of arsenite and
arsenate on amorphous iron hydroxide. Water Res. 16,
1247–1253.
Pirazzoli, P.A., 1996. Sea-level Changes: The Last 20 000
Years. John Wiley, Chichester.
Plumlee, G.S., Smith, K.S., Montour, M.R., Ficklin, W.H.,
Mosier, E.L., 1999. Geologic controls on the composition of
natural waters and mine waters draining diverse mineraldeposit types. In: Environmental Geochemistry of Mineral
Deposits. Part B: Case Studies. Chapter 19, pp. 373–432
(Chapter 19).
Ponnamperuma, F.N., Tianco, E.M., Loy, T., 1967. Redox
equilibria in flooded soils: I. The iron hydroxide systems. Soil
Sci. 103, 371–382.
Quentin, K.E., Winkler, H.A., 1974. Occurrence and determination of inorganic polluting agents. Z. für Bakteriol. und
Hygiene 158, 514–523.
Raven, K.P., Jain, A., Loeppert, R.H., 1998. Arsenite and arsenate adsorption on ferrihydrite: kinetics, equilibrium, and
adsorption envelopes. Environ. Sci. Technol. 32, 344–349.
Reuther, R., 1992. Geochemical mobility of arsenic in a flowthrough water-sediment system. Environ. Technol. 13, 813–823.
Reynolds, J.G., Naylor, D.V., Fendorf, S.E., 1999. Arsenic
sorption in phosphate-amended soils during flooding and
subsequent aeration. Soil Sci. Soc. Am. J. 63, 1149–1156.
Riedel, F.N., Eikmann, T., 1986. Natural occurrence of arsenic
and its compounds in soils and rocks. Wissensch. Umwelt 3–
4, 108–117.
Riedel, G.F., 1993. The annual cycle of arsenic in a temperate
estuary. Estuaries 16, 533–540.
Riedel, G.F., Sanders, J.G., Osman, R.W., 1997. Biogeochemical control on the flux of trace elements from estuarine
sediments: water column oxygen concentrations and benthic
infauna. Estuar. Coast. Shelf Sci. 44, 23–38.
Rittle, K.A., Drever, J.I., Colberg, P.J.S., 1995. Precipitation of
arsenic during bacterial sulfate reduction. Geomicrobiol. J.
13, 1–11.
Robertson, F.N., 1989. Arsenic in groundwater under oxidizing
conditions, south-west United States. Environ. Geochem.
Health 11, 171–185.
Robins, R.G., 1987. Solubility and stability of scorodite, FeAsSO4.2H2O: discussion. Am. Mineral. 72, 842–844.
Robinson, B., Outred, H., Brooks, R., Kirkman, J., 1995. The
distribution and fate of arsenic in the Waikato River System,
North Island, New Zealand. Chem. Spec. Bioavail. 7, 89–96.
Rochette, E.A., Li, G.C., Fendorf, S.E., 1998. Stability of
arsenate minerals in soil under biotically generated reducing
conditions. Soil Sci. Soc. Am. J. 62, 1530–1537.
Sancha, A.M., 1999. Full-scale application of coagulation processes for arsenic removal in Chile: a successful case study.
In: Chappell, W.R., Abernathy, C.O., Calderon, R.L. (Eds.),
Arsenic Exposure and Health Effects. Elsevier, Amsterdam,
pp. 373–378.
Sancha, A.M., Castro, M.L., 2001. Arsenic in Latin America:
occurrence, exposure, health effects and remediation. In:
Chapell, W.R., Abernathy, C.O., Calderon, R.L. (Eds.),
Arsenic Exposure and Health Effects IV. Elsevier, Amsterdam,
pp. 87–96.
Schreiber, M.J., Simo, J., Freiberg, P., 2000. Stratigraphic and
geochemical controls on naturally occurring arsenic in
groundwater, eastern Wisconsin, USA. Hydrogeol. J. 8, 161–
176.
Schroeder, W.H., Dobson, M., Kane, D.M., Johnson, N.D.,
1987. Toxic trace-elements associated with airborne particulate matter—a review. Internat. J. Air Pollut. Control
Hazard. Waste Manag. 37, 1267–1285.
Scott, M.J., Morgan, J.J., 1995. Reactions at oxide surfaces. 1.
Oxidation of As(III) by synthetic birnessite. Environ. Sci.
Technol. 29, 1898–1905.
Scudlark, J.R., Church, T.M., 1988. The atmospheric deposition of arsenic and association with acid precipitation.
Atmos. Environ. 22, 937–943.
Seyler, P., Martin, J.-M., 1989. Biogeochemical processes
affecting arsenic species distribution in a permanently stratified lake. Environ. Sci. Technol. 23, 1258–1263.
Seyler, P., Martin, J.-M., 1990. Distribution of arsenite and
total dissolved arsenic in major French estuaries: dependence
on biogeochemical processes and anthropogenic inputs. Mar.
Chem. 29, 277–294.
Seyler, P., Martin, J.-M., 1991. Arsenic and selenium in a pristine river-estuarine system: the Krka, Yugoslavia. Mar.
Chem. 34, 137–151.
Shacklette, H.T., Boerngen, J.G., Keith, J.R., 1974. Selenium,
fluorine, and arsenic in superficial materials of the conterminous United States. US Geol. Surv., Circ. 692. US
Government Printing Office, Washington DC.
Slomp, C.P., Epping, E.H.G., Helder, W., van Raaphorst, W.,
1996. A key role for iron-bound phosphorus in authigenic
apatite formation in North Atlantic continental platform
sediments. J. Mar. Res. 54, 1179–1205.
Smedley, P.L., 1996. Arsenic in rural groundwater in Ghana. J.
Afr. Earth Sci. 22, 459–470.
Smedley, P.L., Edmunds, W.M., Pelig-Ba, K.B., 1996. Mobility
of arsenic in groundwater in the Obuasi area of Ghana. In:
Appleton, J.D., Fuge, R., McCall, G.J.H. (Eds.), Environmental Geochemistry and Health, Geol. Soc. Spec. Publ.
113. Geological Society, London, pp. 163–181.
Smedley, P.L., Nicolli, H.B., Barros, A.J., Tullio, J.O., 1998.
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
Origin and mobility of arsenic in groundwater from the Pampean Plain, Argentina. In: Arehart, G.B., Hulston, J.R. (Eds.),
Water-Rock Interaction. Proc. 9th Internat. Symp., Taupo,
New Zealand, 1998. Balkema, Rotterdam, pp. 275–278.
Smedley, P.L., Macdonald, D.M.J., Nicolli, H.B., Barros, A.J.,
Tullio, J.O., Pearce, J.M., 2000. Arsenic and other quality
problems in groundwater from northern La Pampa Province,
Argentina. British Geological Survey Technical Report,
WC/99/36. British Geological Survey, Keyworth.
Smedley, P.L., Zhang, M., Zhang, G., Luo, Z., 2001a. Arsenic and
other redox-sensitive elements in groundwater from the Huhhot
Basin, Inner Mongolia. In: Cidu, R. (Ed.), Water-Rock Interaction 2001, Vol. 1. Swets & Zeitlinger, Lisse, pp. 581–584.
Smedley, P.L., Kinniburgh, D.G., Huq, I., Luo, Z., Nicolli,
H.B., 2001b. International perspective on naturally occurring
arsenic problems in groundwater. In: Chappell, W.R.,
Abernathy, C.O., Calderon, R.L. (Eds.), Arsenic Exposure
and Health Effects IV. Elsevier, Amsterdam, pp. 9–25.
Smedley, P.L., Nicolli, H.B., Macdonald, D.M.J., Barros, A.J.,
Tullio, J.O., 2002. Hydrogeochemistry of arsenic and other
inorganic constituents in groundwaters from La Pampa,
Argentina. Appl. Geochem. 17, 259–284.
Smith, A.H., Goycolea, M., Haque, R., Biggs, M.L., 1998.
Marked increase in bladder and lung cancer mortality in a
region of northern Chile due to arsenic in drinking water.
Am. J. Epidemiol. 147, 660–669.
Smith, A.H., Hopenhayn-Rich, C., Bates, M.N., Goeden,
H.M., Hertz-Picciotto, I., Duggan, H.M., Wood, R., Kosnett, M.J., Smith, M.T., 1992. Cancer risks from arsenic in
drinking water. Environ. Health Perspect. 97, 259–267.
Smith, A.H., Lingas, E.O., Rahman, M., 2000. Contamination
of drinking-water by arsenic in Bangladesh: a public health
emergency. Bull. WHO. 78, 1093–1103.
Smith, S.L., Jaffe, P.R., 1998. Modeling the transport and
reaction of trace metals in water-saturated soils and sediments. Water Resour. Res. 34, 3135–3147.
Sohlenius, G., 1996. Mineral magnetic properties of late
Weichselian-Holocene sediments from the northwestern Baltic proper. Boreas 25, 79–88.
Sonderegger, J.L., Ohguchi, T., 1988. Irrigation related arsenic
contamination of a thin, alluvial aquifer, Madison River Valley,
Montana, USA. Environ. Geol. Water Sci. 11, 153–161.
Stewart, F. H., 1963. Data of Geochemistry, 6th ed. Chap. Y.
Marine Evaporites. US Geol. Surv. Prof. Pap. 440-Y.
Stumm, W., Morgan, J.J., 1995. Aquatic Chemistry: Chemical
Equilibria and Rates in Natural Waters. Wiley-Interscience,
New York.
Sullivan, K.A., Aller, R.C., 1996. Diagenetic cycling of arsenic
in Amazon shelf sediments. Geochim. Cosmochim. Acta 60,
1465–1477.
Sun, G.F., Dai, G.J., Li, F.J., Yamauchi, H., Yoshida, T.,
Aikawa, H., 1999. The present situation of chronic arsenism
and research in China. In: Chappell, W.R., Abernathy, C.O.,
Calderon, R.L. (Eds.), Arsenic Exposure and Health Effects.
Elsevier, Amsterdam, pp. 123–126.
Sun, G.F., Pi, J.B., Li, B., Guo, X.Y., Yamauchi, H., Yoshida,
T., 2001. Progresses on researches of endemic arsenism in
China: population at risk, intervention actions, and related
scientific issues. In: Chapell, W.R., Abernathy, C.O.,
Calderon, R.L. (Eds.), Arsenic Exposure and Health Effects
IV. Elsevier, Amsterdam, pp. 79–85.
567
Swedlund, P.J., Webster, J.G., 1998. Arsenic removal from
geothermal bore waters: the effect of monosilicic acid. In:
Arehart, G.B., Hulston, J.R. (Eds.), Water-Rock Interaction.
Proc. 9th Internat. Symp., Taupo, New Zealand, 1998.
Balkema, Rotterdam, pp. 947–950.
Taylor, R.M., 1980. Formation and properties of Fe(II)–
Fe(III) hydroxycarbonate and its possible significance in soil
formation. Clay Minerals 15, 369–382.
Thirunavkukkarasu, O.S., Viraraghavan, T., Subramanian,
K.S., 2001. Removal of arsenic in drinking water by iron
oxide-coated sand and ferrihydrite - Batch studies. Water
Qual. Res. J. Can. 36, 55–70.
Thompson, J.M., Demonge, J. M., 1996. Chemical Analyses of
Hot Springs, Pools, and Geysers from Yellowstone National
Park, Wyoming and Vicinity, 1980–1993. US Geol. Surv.
Open File Rep. 96–68.
Thompson, J.M., Keith, T.E.C., Consul, J.J., 1985. Water
chemistry and mineralogy of Morgan and Growler hot
springs, Lassen KGRA, California. Transactions of the
Geothermal Research Council 9, 357–362.
Thornton, I., Farago, M., 1997. The geochemistry of arsenic.
In: Abernathy, C.O., Calderon, R.L., Chappell, W.R. (Eds.),
Arsenic Exposure and Health Effects. Chapman Hall, London, pp. 1–16.
Torii, M., 1997. Low-temperature oxidation and subsequent
downcore dissolution of magnetite in deep-sea sediments,
ODP Leg 161, Western Mediterranean. J. Geomag. Geoelect.
49, 1233–1245.
Tseng, W.P., Chu, H.M., How, S.W., Fong, J.M., Lin, C.S.,
Yeh, S., 1968. Prevalence of skin cancer in an endemic area
of chronic arsenicism in Taiwan. J. Nat. Cancer Inst. 40,
453–463.
Umitsu, M., 1993. Late Quaternary sedimentary environments
and landforms in the Ganges Delta. Sed. Geol. 83, 177–186.
Ure, A., Berrow, M., 1982. Chapter 3. The elemental constituents of soils. In: Bowen, H.J.M. (Ed.), Environmental
Chemistry. Royal Society of Chemistry, London, pp. 94–203.
Varsányi, I., Fodré, Z., Bartha, A., 1991. Arsenic in drinking
water and mortality in the southern Great Plain, Hungary.
Environ. Geochem. Health 13, 14–22.
Villalobos, M., Trotz, M.A., Leckie, J.O., 2001. Surface complexation modeling of carbonate effects on the adsorption of
Cr(VI), Pb(II), and U(VI) on goethite. Environ. Sci. Technol.
35, 3849–3856.
Wang, G., 1984. Arsenic poisoning from drinking water in
Xinjiang. Chin. J. Prevent. Med. 18, 105–107.
Wang, L., Huang, J., 1994. Chronic arsenism from drinking
water in some areas of Xinjiang, China. In: Nriagu, J.O.
(Ed.), Arsenic in the Environment, Part II: Human Health
and Ecosystem Effects. John Wiley, New York, pp. 159–172.
Waslenchuk, D.G., 1979. The geochemical controls on arsenic
concentrations in southeastern United States rivers. Chem.
Geol. 24, 315–325.
Waychunas, G.A., Fuller, C.C., Rea, B.A., Davis, J.A., 1996.
Wide angle X-ray scattering, WAXS study of ‘‘two-line’’
ferrihydrite structure: Effect of arsenate sorption and counterion variation and comparison with EXAFS results. Geochim. Cosmochim. Acta 60, 1765–1781.
Webster, J.G., 1999. Arsenic. In: Marshall, C.P., Fairbridge,
R.W. (Eds.), Encyclopaedia of Geochemistry. Chapman
Hall, London, pp. 21–22.
568
P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
Webster, J.G., Nordstrom, D.K., Smith, K.S., 1994. Transport
and natural attenuation of Cu, Zn, As, and Fe in the acid
mine drainage of Leviathan and Bryant Creeks. In: Alpers,
C.N., Blowes, D.W. (Eds.), Environmental Geochemistry of
Sulphide Oxidation. Am. Chem. Soc. Symp. Ser., 550. ACS,
Washington DC, pp. 244–260.
Wegelin, M., Gechter, D., Hug, S., Mahmud, A., 2000.
SORAS—a simple arsenic removal process. In: ‘Water, Sanitation, Hygiene: Challenges of the Millennium’, 26th WEDC
Conference, Dhaka, 5–9 November, 2000, pp. 379–382.
Welch, A.H., Lico, M.S., 1998. Factors controlling As and U in
shallow ground water, southern Carson Desert, Nevada.
Appl. Geochem. 13, 521–539.
Welch, A.H., Helsel, D.R., Focazio, M.J., Watkins, S.A., 1999.
Arsenic in ground water supplies of the United States. In:
Chappell, W.R., Abernathy, C.O., Calderon, R.L. (Eds.),
Arsenic Exposure and Health Effects. Elsevier, Amsterdam,
pp. 9–17.
Welch, A.H., Lico, M.S., Hughes, J.L., 1988. Arsenic in
ground-water of the Western United States. Ground Water
26, 333–347.
Welch, A.H., Westjohn, D.B., Helsel, D.R., Wanty, R.B., 2000.
Arsenic in ground water of the United States: occurrence and
geochemistry. Ground Water 38, 589–604.
Wersin, P., Hohener, P., Giovanoli, R., Stumm, W., 1991.
Early diagenetic influences on iron transformations in a fresh
water lake sediment. Chem. Geol. 90, 233–252.
White, D.E., Hem, J.D., Waring, G.A., 1963. Data of Geochemistry, 6th ed. M. Fleischer, (Ed). Chapter F. Chemical Composition of Sub-Surface Waters. US Geol. Surv. Prof. Pap. 440-F.
WHO, 1993. Guidelines for drinking-water quality. Volume 1:
Recommendations, 2nd ed. WHO, Geneva.
WHO, 2001. Environmental Health Criteria 224: Arsenic compounds 2nd edition. World Health Organisation, Geneva.
Widerlund, A., Ingri, J., 1995. Early diagenesis of arsenic in
sediments of the Kalix River estuary, Northern Sweden.
Chem. Geol. 125, 185–196.
Wilkie, J.A., Hering, J.G., 1996. Adsorption of arsenic onto
hydrous ferric oxide: Effects of adsorbate/adsorbent ratios
and co-occurring solutes. Colloids Surfaces A-Physicochem.
Engineer. Aspects 107, 97–110.
Wilkie, J.A., Hering, J.G., 1998. Rapid oxidation of geothermal arsenic(III) in streamwaters of the eastern Sierra
Nevada. Environ. Sci. Technol. 32, 657–662.
Williams, M., 1997. Mining-related Arsenic Hazards: Thailand
Case-study. Summary Report. Brit. Geol. Surv. Tech. Rep.,
WC/97/49.
Williams, M., Fordyce, F., Paijitprapapon, A., Charoenchaisri,
P., 1996. Arsenic contamination in surface drainage and
groundwater in part of the southeast Asian tin belt, Nakhon
Si Thammarat Province, southern Thailand. Environ. Geol.
27, 16–33.
Wilson, F.H., Hawkins, D.B., 1978. Arsenic in streams, stream
sediments and ground water, Fairbanks area, Alaska.
Environ. Geol. 2, 195–202.
Wyatt, C.J., Fimbres, C., Romo, L., Mendez, R.O., Grijalva,
M., 1998. Incidence of heavy metal contamination in water
supplies in Northern Mexico. Environ. Res. 76, 114–119.
Xu, H., Allard, B., Grimvall, A., 1988. Influence of pH and
organic substance on the adsorption of As(V) on geologic
materials. Water Air Soil Pollut. 40, 293–305.
Xu, H., Allard, B., Grimvall, A., 1991. Effects of acidification
and natural organic materials on the mobility of arsenic in
the environment. Water Air Soil Pollut. 57–8, 269–278.
Yan, X.-P., Kerrich, R., Hendry, M.J., 2000. Distribution of
arsenic(III), arsenic(V) and total inorganic arsenic in porewaters from a thick till and clay-rich aquitard sequence,
Saskatchewan, Canada. Geochim. Cosmochim. Acta 64,
2637–2648.
Yokoyama, T., Takahashi, Y., Tarutani, T., 1993. Simultaneous determination of arsenic and arsenious acids in geothermal water. Chem. Geol. 103, 103–111.
Yusof, A.M., Ikhsan, Z.B., Wood, A.K.H., 1994. The speciation of arsenic in seawater and marine species. J. Radioanal.
Nucl. Chem.-Articles 179, 277–283.
Zaldivar, R., 1974. Arsenic contamination of drinking water
and foodstuffs causing endemic chronic poisoning. Beitr. zur
Pathol. 151, 384–400.
Zhai, C., Dai, G., Zhang, Z., Gao, H., Li, G., 1998. An environmental epidemiological study of endemic arsenic poisoning in Inner Mongolia. In: Abstr. 3rd Internat. Conf. Arsenic
Exposure and Health Effects, San Diego, 1998, p. 17.
Zhu, B.J., Tabatabai, M.A., 1995. An alkaline oxidation
method for determination of total arsenic and selenium in
sewage sludges. J. Environ. Qual. 24, 622–626.
Zobrist, J., Dowdle, P.R., Davis, J.A., Oremland, R.S., 2000.
Mobilization of arsenite by dissimilatory reduction of adsorbed arsenate. Environ. Sci. Technol. 34, 4747–4753.
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